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BIOGENIC EMISSIONS TO THE ATMOSPHERE _ . . A wide variety of natural products, including volatile substances and particulates, are emitted into and removed from the atmosphere by both terrestrial and aquatic systems. Terrestrial reactions are better known, owing to the large body of agricultural research, and knowledge about the compounds thus generated or removed is increasing constantly. Assessments will need to be modified frequently as information is obtained on factors affecting rates of transformation and on local, regional, and global emissions and removals. Soils emit a variety of volatile products and also remove volatile substances from the gas phase. The activities of microorganisms dominate the soil ecosystem, and most evidence indicates that heterotrophic microorganisms are the chief agents for the production of many gases. The biochemical mechanisms involved in the microbial generation of these gases vary markedly, however, and no generalization covers all of the volatile products or even any significant class of volatiles. NITROGEN Terrestrial Ecosystems Ecosystems consume atmospheric molecular nitrogen (N2) through biological nitrogen fixation and also evolve it as one of the products of denitrification. Biological fixation of N2 leads to incorporation of nitrogen compounds into soil. The annual global fixation of molecular nitrogen by the land mass has been estimated by a number of authors. One such estimate indicates that the annual amount of molecular nitrogen fixed into soil globally is 99 x 1012 grams of nitrogen per year (Delwiche and Likens 1977). Denitrification is the process by which some microorganisms can reduce nitrate (NO3) or nitrite (NO2-) to the gaseous forms molecular nitrogen (N2) and nitrous oxide (N2O), which may then be lost to the atmosphere. This is a major pathway for the loss of fixed nitrogen from an ecosystem. An enormous amount of work is being done 22
23 on denitrification, and the topic has been reviewed recently (Delwiche and Bryan 1976, Delwiche et al. 1978~. Some estimates of denitrification have been made by measuring the quantity of nitrous oxide generated. Most of these studies have been conducted in the laboratory, but a number of studies under field conditions (Focht and Stolzy 1978, Denmead 1979) have shown that nitrous oxide generation may occur in both dry-land agriculture and in flooded fields (Denmead et al. 1979~. The rate of denitrification varies extensively with season of the year and oxygen content of the soil (Dowdell and Smith 1974~. The major factors that affect nitrous oxide emissions are nitrate content of the soil, oxygen status, moisture, pH, and temperature. It is generally believed that nitrous oxide is formed by the microbial reduction of nitrate or nitrite as microorganisms use these compounds for the terminal electron acceptor in their metabolism. An intriguing recent investigation, however, suggests that some nitrous oxide may also be generated in soils during Vitrification, the microbial oxidation of ammonia to nitrite and nitrate (Bremner and Blackmer 1978, Freney et al. 19781. This formation of nitrous oxide from ammonia was earlier shown to take place in cultures of autotrophic Vitrifying bacteria (Yoshida and Alexander 1970~. Nitric oxide (NO), too, is also produced in soil. Presumably, this gas is generated in denitrification as nitrate is reduced. Nitrate is the dominant if not the sole precursor of nitric oxide. The quantity of nitric oxide formed is markedly affected by temperature, pH, and moisture (Bailey and Beauchamp 1973, Garcia 1976), and much of the release is apparently microbial (Garcia 1976~. Calculations have been made of the total quantity of inorganic nitrogen volatilized through denitrification. One estimate Is 120 x 1012 grams of nitrogen per year (Delwiche and Likens 1977~. An early estimate of the emission of nitrogen oxides gave values for biological release globally of 120 x 1012 grams nitrogen per year of nitrous oxide and a figure of 234 x 1012 grams nitrogen per year of nitric oxide (Robinson and Robbins 1970a). A more recent estimate suggests that the quantity of nitrogen oxides emitted from natural sources on the earth's surface in the Northern Hemisphere is probably no more than 30 x 1012 grams of nitrogen per year (Galbally 1976~. It is obvious from the range of these estimates that the data base is limited, and the conceptual models are crude. In making such measurements, it is difficult to separate evolution and uptake of nitrogenous gases. Nitrous oxide is both emitted and removed from the gas phase. There is some evidence that net removal takes place largely in waterlogged soils (Freney et al. 1978~. Denitrification is promoted by anaerobiosis and by the presence of organic materials that stimulate microbial proliferation. Altogether, soils are a more significant source than a sink for nitrous oxide (Blackmer and Bremner 1976~. Agricultural scientists and technologists have extensively studied ammonia losses from soil. Most of these studies reflect anthropogenic inputs of nitrogen because they have been done on fertilized, tilled soils or pastured areas. Many studies were prompted by concern with
24 the loss of the nitrogen applied to the soil in the form of fertilizer for crop production. In some instances, losses can be quite appreciable; for example, in one study, it was noted that the ammonia-nitrogen loss varied from 19 to 50 percent of the quantity of fertilizer nitrogen applied, the extent of loss being dependent on the fertilization rate and the temperature (Fenn and Kissel 1974). Despite this large body of information, the inputs of ammonia to the atmosphere from large land areas have not been measured directly except in a few instances. In one such study, the flux of ammonia over a grazed pasture in Australia was determined for a 3-week period in late summer. The quantity of ammonia evolved was equivalent to 260 grams per hectare per day (10.8 g/ha/hr), a quantity that is a substantial part of the nitrogen cycled in the pasture. It was estimated that, under these conditions, a major source of the ammonia was probably the urine of the sheep in the pasture (Denmead et al. 1974~. Other studies have indicated that the losses in Australian pastures may be up to 13 grams of nitrogen per hectare per hour when the pasture is grazed but only about 2 grams of nitrogen per hectare per hour when it is not grazed (Denmead et al. 1976). A number of investigations have verified the significance of animal manure, particularly from dairy cattle, to the volatilization of ammonia and its presence in the overlying air (Luebs et al. 1973, Lauer et al. 1976). Data from the U.S. National Atmospheric Deposition Program (Gibson and Baker 1979) show large amounts of ammonia in precipitation in areas of the midwest where feedlots are abundant and where anhydrous ammonia is a popular agricultural fertilizer. Global estimates, based upon unfortunately few data, have been made by several individuals. The consensus has been that microbial action in soil is a major source of ammonia generated from the land surface but the size of estimates varies greatly. One estimate of the ammonia loss from the soil is 960 x 1012 grams per year (Robinson and Robbins 1970a); another investigator has suggested that the natural emissions of ammonia in the Northern Hemisphere do not exceed 130 x 1012 grams of ammonia nitrogen per year (Galbally 1976). Several organic nitrogen compounds, including amines and nitrogen heterocycles, may also be evolved from soils. Research has focused attention on the generation of these products in lands treated with animal manure, either from cattle or poultry. Among the amines formed are mono-, di- and trimethylamine, ethylamine, other alkylamines, indole, and skatole. These products are probably the result of microbial activity in decomposing animal wastes, and the process is probably also occurring under the anaerobic conditions that prevail in heaps of these organic materials (Young et al. 1971, White et al. 1971, Rouston et al. 1977, Mbsier et al. 1973~. Volatile organic nitrogen, presumably as amines, may be detected not only immediately above the decaying organic matter but also in traps placed at sites adjacent to high densities of animals (Elliott et al. 1971~.
25 Aquatic Ecosystems Much information on nitrogen in aquatic ecosystems is summarized in a recent National Researach Council report (1978d). We shall therefore review only new or especially relevant aspects here. Nitrogen fixation in aquatic ecosystems Is carried out largely by a few genera of blue-green algae (Brezonik 1977, Jones and Steward 1969~. These genera become abundant in lakes where the supply of nitrogen is low relative to that of phosphorus (Schindler 1977~. A survey of recent literature revealed that atmospheric molecular nitrogen is not fixed in lakes that have inputs of nitrogen 10 times greater by weight than phosphorus (Flett et al. 1980~. Nor did this survey find fixation of N2 in either oxic or anoxic lake sediments, although nitrogen was fixed by the periphyton community of a small lake with a very low input ratio of N to P. In two small lakes, one studied for three years, nitrogen fixation supplied 7 to 38 percent of the total annual income of nitrogen. Rates of nitrogen fixation depended on light intensity, much as carbon fixation does. It is impossible to make large-scale estimates of N2 fixation in fresh water, because there are few quantitative studies for large water bodies. Howard et al. (1970) showed that algal fixation occurred in Lake Erie but supplied no quantitative estimates. Vanderhoef et al. (1974) showed that fixation supplied only a small part of the nitrogen input to Green Bay, a highly eutrophic part of the Laurentian Great Lakes. This makes it seem unlikely that most other large lakes, which are more oligotrophic, fix substantial amounts of N2, and lakes are probably minor participants in the global nitrogen cycle. A technical problem renders uncertain many quantitative estimates of nitrogen fixation in either terrestrial or aquatic systems. Most studies to date have employed the convenient, sensitive, and inexpensive acetylene-reduction method, which assumes that every three moles of acetylene reduced to ethylene by the nitrogenase system represent the equivalent of one mole of molecular nitrogen fixed (Graham et al. 1980~. In two comparisons, however, the actual conversion ratio has varied between 1.7 and 9.1 moles of acetylene reduced per mole of nitrogen fixed (Hardy et al. 1968, Graham et al. 1980~. Because of this variability in the conversion ratio estimates of total nitrogen fixed based solely on the acetylene method may be in error by a factor of two to four. Denitrification rates are highest in eutrophic lakes, where the combination of a rich organic substrate, high nitrate and low oxygen favors the denitrifying microbiota (Brezonik 1977~. The epilimnion sediment-water interface appears to be the most important site (Tiren et al. 1976, Chan and Campbell 1980), although the thermocline region can also be important when sufficient nitrate is present along with low oxygen (< 0.2 mg/1~. Denitrification can also take place in anoxic hypolimnia, with nitrous oxide often occurring as an end product. The importance of denitrification at this site is limited, because high nitrate concentrations usually do not occur under anoxic conditions.
26 Chan and Campbell (1980) estimate that only 1.4 percent of the nitrate entering a small eutrophic Canadian lake was denitrified. They found that a combination of nitrate-nitrogen (>0.01 mg/1) and oxygen (< 0.2 mg/1) was necessary for significant denitrification to take place. In coastal and marine sediments there is some evidence that ammonia may be a significant end product of nitrate reduction (Koike and Hattori 1978, Sorensen 1978~. Regional data for aquatic ecosystems are too few to warrant global estimates. SULFUR Terrestrial Ecosystems It is generally accepted that soils are sources of volatile sulfur compounds. Estimates have given widely dissimilar figures; emissions from the land surface, where the reaction is generally believed to be biological, for example, have been estimated at 68 x 1012 (Robinson and Robbins 1970b), 58 x 1012 (Friend 1973), and as little as 3 x 1012 grams of sulfur per year (Bolin and Charlson 1976~. The authors of most of the estimates believe that the sulfur emitted biologically from soil is derived from vegetation decaying in intertidal flats, swamps, and bogs, and it has generally been assumed that the product evolved was hydrogen sulfide. But monitoring studies and direct analysis of soils indicate that any hydrogen sulfide formed would probably react with iron compounds to form iron sulfide (Bloomfield 1969~. Volatile sulfur compounds are formed in soil samples treated with sulfur-containing amino acids. Such soils evolve methyl mercaptan, clime thyl sulfide, dimethyl disulfide, ethyl mercaptan, ethyl methyl sulfide, diethyl disulfide, carbon disulfide, and small amounts of carbonyl sulfide. Some of these products are detected during the decomposition of animal manure {Banwart and Bremner 1975), and soils treated with a variety of natural materials, including sewage sludge and plant remains, evolve some of the same gases, but no hydrogen sulfide is emitted. The conversions are affected by the dominant amino acids present and by the level of oxygen (Banwart and Bremner 1976a, b). Hence, the suggestion that hydrogen sulfide emissions are major sources of the sulfur coming from soil is apparently incorrect. No large areas of the globe that evolve hydrogen sulfide have been discovered. The chief product coming from adequately aerated soil is more likely to be dimethyl sulfide (Rasmussen 1974). There are only limited measurements of the release of volatile biogenic sulfur compounds from soil (Adams and Farwell 1980, Aneja et al. 1980, A. Goldberg et al. 1981), and therefore the global significance of soil as a source is difficult to evaluate at this time.
27 Aquatic Ecosystems The reduction of sulfate to sulfide occurs in fresh waters, marine areas, and fresh and salt marshes under anoxic conditions. The bacterial genus Desulfovibrio is responsible, and rates are apparently highest in tidal salt marshes, where tides replenish the supply of sulfates. In such salt marshes, sulfate reducers actually contribute as much energy to the salt marsh food chain as direct photosynthetic fixation. Howarth and Teal (1979) report an annual rate of 75 mol SO4 per square meter for a New England marsh. Highest rates occurred in autumn, when decaying plants increased the available organic substrate. In fresh waters, production appears to be largely in anoxic hypolimnia or in sediments rich in organic matter. Measurements are too few to allow reliable regional or global estimates. The rate of reduction is dependent on sulfate concentrations, as well as on the available organic matter, both of which vary widely in different lakes. The amount of sulfide that actually reaches the atmosphere under natural conditions is not known. AS in terrestrial soils, iron concentrations in the active areas are often high enough to exceed the solubility product of ferrous iron and sulfide, and in these cases the element is precipitated as iron monosulfide or pyrite (Howarth 1979, Schindler et al. 1980a, Cook 1981~. In salt marshes, this precipitation is thought to prevent the buildup of high concentrations of soluble sulfide, which could be toxic to marsh grasses and other microbiota (Howarth and Teal 1979~. Much of the iron sulfide is reoxidized as part of the seasonal cycle. Too few ecosystems have been investigated to permit assessment of how widespread the above phenomena may be. Both lakes and marshes may vary greatly in their iron content, depending on underlying geology. In iron-impoverished areas, substantial volatile sulfide may be released to the atmosphere. Both marine and freshwater plants appear to produce clime thyl sulfide and dimethyl sulfoxide as metabolic by-products. When emitted to the atmosphere, these methylated sulfur compounds are rapidly oxidized to sulfur dioxide or sulfuric acid. The subject is reviewed by Andreae (19801. Lovelock et al. (1972) measured dimethylsulfide and carbonyl sulfide in air over the Atlantic Ocean, and assumed the compounds to be of marine origin. Subsequently, Nguyen et al. (1978) measured dimethylsulfide in sea water and calculated that the oceans could contribute substantially to the natural global sulfur budget. Maroulis and Brandy (1977), however, estimated dimethylsulfide emissions from Atlantic coastal regions of the United States to be less than 6 milligrams per square meter per year, which would make an inconsequential contribution to the natural global sulfur budget. Thus, total natural sulfur emissions from aquatic ecosystems and wetlands cannot be quantified owing to the paucity of studies. Likewise, the small number and extreme variability of estimates of volcanic emissions of sulfur dioxide makes it extremely difficult to quantify total natural emissions of sulfur dioxide to the atmosphere. Available estimates of sulfur dioxide emissions for single volcanoes vary by a factor of 10 (Hoff and Gallant 1980~.
28 TRACE METALS There are several lines of evidence that terrestrial plants release metals to the atmosphere. For example, pea plants grown in solutions containing radioactive zinc released the metal to the atmosphere (Beauford et al. 1975~. The rate of zinc mobilization was about 1 microgram/h/m2 of leaf. Any extrapolation of these results to environmental situations is unwarranted; still, the data do suggest that there may be an appreciable flow of zinc from plants into the atmosphere. Supporting evidence comes from field studies of plant exudates (Cur tin et al. 1974~. Polyethylene bags were placed around pine and fir trees, and the exudates, condensed in these bags over period of five to ten days, were collected and analyzed. The ash of the residue of volatile exudates contained lithium, beryllium, boron, sodium, magnesium' titanium, vanadium, chromium' manganese, iron, cobalt' nickel, copper r zinc, gallium, arsenic, strontium' yttrium, zirconium' molybdenum, silver J lead, bismuth' cadmium, tin, antimony, barium, and lanthanum. The underlined elements were most markedly enriched in the exudates compared to the ash of the associated vegetation. Curtin and colleagues attributed the volatilization of some of the metals to completing with terpenes and urged the initiation of an air sampling program to assess this possibility. Biological methylation on both land and sea may result in the transfer of metals from the earth's surface to the atmosphere. It is believed that mercury can enter the air from the sea surface in the form of gaseous dimethyl mercury (Wood and Goldberg 1977, Tomlinson et al. 1980~. In the atmosphere it may disproportionate into ethane, methane, and mercury. Methylated forms of mercury are also produced in sediments, and diffusion into the overlying water may lead to these substances entering the atmosphere. Methylmercury formation is thought to be almost solely the result of microbiological processes (Woolson 1977~. Various mercuric compounds can be used as substrates, including inorganic compounds, elemental metal, and acetate (Rogers 1979, Hamilton 1972~. Several investigators have demonstrated that mercury in soil can be converted to volatile products. The chief if not the sole agents of this volatilization are microorganisms, because the process is reduced markedly or abolished by sterilization of the coil - McFarlane 1979~. AS in aquatic systems, the conversion of the element from a nonvolatile to a volatile state occurs With a number of forms of inorganic mercury as well as with mercuric acetate (Rogers 1979~. One of the products, has been identified as methylmercury, the production of which is affected by soil moisture, temperature, and mercury concentration. High soil temperatures appear to be favorable to the production of the volatile derivative, but the conversion is retarded by excessive moisture or dry conditions (Rogers 1976, Landa 1979~. Elemental mercury is also a significant source of mercury emissions from soil. Most of the mercury in the atmosphere appears to be in the elemental form (William Fitzgerald, Department of Chemistry, University of Connecticut, Storrs, personal communication). The ~ (Rogers and As in aquatic systems, the conversion of the with a number of forms
29 clime thyl mercury emitted into the atmosphere later decomposes to elemental mercury. These investigations with mercury have prompted studies with other trace metals that can be methylated and volatilized from plant surfaces or internal plant tissues, such as arsenic, selenium, tin, germanium, and lead (e.g., Beauford et al. 1975, Jernelov and Martin 1975~. Concern has been expressed about the possible toxic effects of the lipid-soluble metalalkyls of the above metals upon the central nervous system of higher organisms. Methylated tin compounds have been measured in sea waters and algae from the coast of California. Dimethyl and tetramethyl tin compounds have been found in plants at concentrations usually exceeding those of inorganic tin (Hodge et al. 19791. In laboratory experiments, moreover, the protonated tin compound, stannane (SnH4), is produced during the decomposition of plants. These tin compounds are volatile enough to be introduced from surface waters to the atmosphere where they are easily oxidized. It should be noted that human activities appear to have increased the tin levels in marine and fresh-water systems by about an order of magnitude (Seidel et al. 1980~. There is a large enrichment of selenium in the atmosphere, and Duce et al. (1974) have suggested that this selenium is evolved by natural rather than anthropogenic sources. By analogy to the microbial transformation of sulfur, which is similar to selenium, one might expect that the product would be a methyl selenium derivative. Indeed, studies of a number of individual microorganisms indicate that they can convert selenate, selenite, elemental selenium, and organic selenium compounds to dimethylselenide. One of the organisms studied can also produce hydrogen selenide (Doran and Alexander 1977a). Studies under laboratory conditions indicate that selenium may be volatilized when soils are amended with elemental selenium, selenite, selenate, and some organic selenium compounds. These emissions apparently are the result largely or entirely of the activity of microorganisms, and the processes are promoted if organic materials are added to the soil. The chief product appears to be dimethyl- selenide, although clime thyl diselenide may be formed from certain organic sulfur compounds, and hydrogen selenide is possibly emitted as well (Doran and Alexander 1977b). Microorganisms have been obtained in axenic cultures that can use certain volatile organic selenium compounds as carbon sources for growth (Doran and Alexander 1977a). The significance of this in vitro metabolism of organic selenium compounds to natural conditions and the significance of the products of these reactions to selenium levels in the atmosphere remain unknown. According to Lewis (1976), selenium is released from the leaves of selenium-accumulating Astragalus sp. (locoweeds) as organo-selenium compounds, including methyl- and dimethylselenides. Inquiries into problems associated with human poisoning at the turn of the century determined that arsenic-containing wallpaper components were being converted by microorganisms into volatile arsenic derivatives (Gosio 1897~. More recently, it has been established that microorganisms are able to convert a number of
30 arsenic compounds, both inorganic and organic, to volatile alkyl arsenic derivatives (Cox and Alexander 1973~. Soils treated with arsenicals emit volatile products. In the early investigations of flooded soils, volatile arsenic products were not identified (Epps and Sturgis 1939). Lately, the products have been identified as methylarsenic derivatives, and the recent studies point to the possibility of soils thus transferring arsenic to the overlying atmosphere (Cheng and Focht 1979, Woolson 1977~. Figure 3.1 summarizes the biological cycles of two trace metals, mercury and arsenic. Microorganisms are known to metabolize Hg, As, Sn, Se, and S and form methylated compounds of these metals. A number of organisms can methylate tellurium (Fleming and Alexander 1972, Challenger 1951), and methylation reactions characterize the behavior of certain microorganisms when exposed to a number of toxic elements. ATMOSPHERIC EMISSIONS FROM BURNING OF NATURAL BIOMASS Supplementing such low-temperature biological processes is the burning--both natural and man-induced--of plants, which can introduce substantial quantities of metals as well as carbon, nitrogen, and sulfur compounds into the atmosphere. The exact amounts are not known, but forest fires often create temperatures of 400°C and up, clearly sufficient to vaporize some inorganic compounds. In the United States alone, forest fires have been estimated to contribute annually 27 megatons (2.4 x 1013g) of particulates to the atmosphere and an additional 17 megatons (1.5 x 1013g) per year are estimated to originate from slash burning and forest litter control operations (Robinson and Robbins 1971). There is in general a latitudinal zonation of forest types about the earth's surface. Tropical forests span the equatorial regions, but the "combustion" of their organic matter is usually through low-temperature, biologically mediated processes. The global contribution from forest fires of particles in the atmosphere is estimated to be 150 megatons (1.4 x 1014g) per year. Recent studies have raised interesting new questions concerning the influence that burning of biomass--wood and other natural products excluding fossil fuels--may have on global atmospheric budgets of certain trace gases such as CO2, CO, H2 J N2OJ NO, NO2 t and COS (Adams et al. 1977; Wong 1978; Radke et al. 1978; Crutzen et al. 1979; Seller and Crutzen 1980). These studies have used very limited, preliminary data on emission factors and total biomass burned to calculate regional and global inputs from burning. Estimates of atmospheric input of CO2 from burning have been particularly controversial because of the extremely important climate implications. Field studies currently in progress may improve our understanding of emissions from large-scale intentional burning. The growing popularity of wood as a fuel for domestic heating and certain industrial heat needs has also stimulated considerable effort to understand environmental effects on air quality (Cooper 1980; Hall and DeAngelis 1980). A summary of some measured emissions of major
31 Mercury Arsenic Air CH4 + - / Water / CH3Hg+ \ ~ C2H. Hg° (CH3)2Hg HR°WCH Hg+ Bacteria (CH ) Hg R~tPri~ '\ Bacteria\ /acteria g2+ H~ Hg° Bacte ria Sediment Air CH3 HO As. CH3 ,~ 1 1 %\ \ O ~ \ ~ [02]~/ \ l \ Water C l 3 C l 3 CH3- As3 CH3 + H As3 CH3 Trimethyla rsi n e Dimeth yla rsi ne \ \ \ ~ Mol~~~3acteria ~ | OH \. CH~ CH3 HO As! OH: ~ As! OH :* HO As! OH ~ ~ HO As+ CH3 i Bacteria l l Bacteria l l Bacteria O O o o Arsenate Arsenite Methylarsenic Dimethylarsinic Sediment acid acid FIGURE 3.1 The biological cycle for mercury and arsenic. SOURCE: Wood (1974). Reprinted with permission from Science 183:1049-1052. Copyright ~ 1974 by the American Association for the Advancement of Science.
32 pollutants from residential wood combustion sources is given in Table 3.1. There is preliminary evidence indicating that biomass burning is a source of trace metals and potentially carcinogenic trace organic compounds (Strum and Loveland 1974; Bumb et al. 1980; Cooper 1980). The composition of emissions from biomass burning of any type depends on a number of variables including the temperature of combustion, biomass physical and chemical properties, and location of the burn (i.e., forest, range, or wood burner, etc.). Considerable research will be required to provide an adequate data base on emission factors, transport, and fate for combustion products from biomass burning . NATURAL ORGANIC PRODUCTS Little is known about the amounts and types of organic compounds emitted naturally from the biosphere to the atmosphere. The limited information available has been reviewed recently (Duce 1978, NSF 1979) and is summarized in Table 3.2. It is generally accepted that most natural hydrocarbons emitted to the atmosphere in gaseous form are of low molecular weight--for example, isoprene and terpene emitted from terrestrial vegetation and methane generated in bog and lake sediments. The total annual emissions from the entire biosphere are estimated to be 850 x 1012 g of carbon. In addition to emissions from vegetation, the ocean and natural fires are known to contribute small amounts of hydrocarbons. No estimates are available for total emissions from soils, fresh water s, or wetlands. Likewise, large uncertainties are associated with estimates of natural emissions of organic particulates. Natural fires are known to contribute, but there is no quantitative information for releases from oceans, terrestrial vegetation, soils, fresh waters, or wetlands. The organic products of burning processes have been studied primarily in sediments, both lacustrine and marine (D.M. Smith et al. 1973, Muller et al. 1977, Laflamme and Hites 1978, Windsor and Hites 1979, Blumer et al. 1977, Blumer and You ng blood 1975). The organic compounds formed include anthracenes, phenanthrenes, pyrenes, fluoranthenes, chrysenes, triphenylenes, benzanthracenes, and others, as well as elemental carbon. SUMMARY A number of volatile products are generated from soils, vegetation or water surfaces both under natural conditions and after stimulation of the biota by anthropogenic activities. These volatile products include compounds of nitrogen, sulfur, numerous heavy metals, and organic substances. Both low-temperature, microbially mediated processes and high-temperature volatilization, as in forest fires, are responsible. The release of aerosols may be important in many cases. In all cases, data are too few to allow meaningful interpretation. Global estimates that have been made for emitted substances are little better than order-of-magnitude values.
33 TABLE 3.1 Emissions of Major Pollutants from Residential Wood Combustion Wood-Burn~ng Stoves Fireplaces Grams per Pounds Percent- Grams per Pounds Percent- Kilogram per 106 age Parti- Kilogram per 106 ageParti- Chem~cal Species of Wood Btu curates of Wood Btu curates Carbon monoxide 160 22 22 3.0 (83-370) (1 1-40) Volatile hydra- 2.0 0.28 19 2.6 carbons (0.3-3.0) NOxasNO2 0.5 0.07 1.8 0.25 SOxas SO2 0.2 0.03 Aldehydes 1.1 0.15 1.3 0.18 Condensable 4.9 0.67 58 6.7 0.92 organics (2.2-14) (5.4-9.1) Particulates 3.6 0.50 42 2.4 0.33 (0.6-8.1) (1.8-2.9) Total particulates 8.5 1.2 100 9.1 1.3 (1-24) (7.2-1 2) Polycyclic organic material Benzo(a)pyrene Carcinogensa Priority pollutantsb Na Al s S C1 K Ca Organic carbon Elemental carbon 0.3 0.04 0.0025 0.0003 0.038 0.005 0.41 0.06 0.005 0.0007 0.004 0.0006 0.003 0.0004 0.03 0.004 0.05 0.007 0.07 0.01 0.004 4.2 0.7 0.0006 0.58 0.1 74 26 100 3.5 0.03 0.00073 0.45 0.0059 4.8 0.063 0.06 0.004 0.05 0.002 0.04 0.002 0.4 0.004 0.6 0.05 0.8 0.05 0.05 0.005 49 4.2 8 1.2 0.004 0.3 0.0001 0.008 0.0008 0.06 0.009 0.7 0.0006 0.04 0.0003 0.02 0.0003 0.02 0.0006 0.04 0.007 0.6 0.007 0.5 0.0007 0.05 0.58 46 0.16 13 aIncludes benz (a)anthracene, diben zanthracene, benzo (c)phenanthrene, benzofluor- anthenes, methylcholanthene, benzopyrenes, dibenzopyrenes, and dibenzocarbonzoles. bIncludes acenaphthylene, fluorene, anthracene/phenanthrene, phenol, fluoranthene, pyrene, benz(a)anl:hracene, benzofluoranthenes, benzo(a)pyrene, benzo(ghi)perylene, dibenzanthracenes, acenaphthene, and ethyl benzene. SOURCE: Cooper (1980). Cooper used values from a variety of sources discussed in his paper. Reprinted, with permission, from Journal of the Air Pollution Control Association 30:85 5-86 1.
34 TABLE 3.2 Natural Sources of Organic Carbon (X 10~2 g of carbon per year) Gaseous Nonmethane Particulate Source Hydrocarbons Organics Vegetation Isoprene 350 ? Terpenes 480 Others ? Total >8 30 Soil ? 1 1a Ocean/freshwater 1.7 14 Biomass burning 3.5-46 1.6-7.1 Total ~830-880 ~15-21a aSoil-derived organics are mostly associated with large particles and are not included in the global budget. SOURCE: Adapted from Duce (1978) and National Science Foundation (1979). .
