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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
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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
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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~.
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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.
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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.
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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~.
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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
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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
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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
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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.
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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.
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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.
OCR for page 34
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).
.
Representative terms from entire chapter:
molecular nitrogen
