GLOBAL BIOGEOCHEMICAL CYCLES, VOL. 6, NO. 4, PAGES 459-483, DECEMBER 1992 

GASEOUS AMMONIA FLUXES AND BACKGROUND 

CONCENTRATIONS IN TERRESlRIAL ECOSYSTEMS 

OF THE UNITED STATES 

A. 0. Langford I 

Cooperative Institute for Research in Environmental Science, 

University of Colorado, Boulder 

F. C. Fehsenfeld2 

Aeronomy Laboratory, National Oceanic and Atmospheric 

Administration, Boulder, Colorado 

J. Zachariassen3 and D. S. Schimel4 

Natural Resources Ecology Laboratory, Colorado State 

University, Fort Collins 

Abstract. Ammonia (NH3) is the dominant gaseous base 

in the atmosphere and the principal neutralizing agent for 

atmospheric acids, yet remains one of the least well 

characterized atmospheric trace compounds. In particular, the 

spatial and temporal distribution of the background 

concentrations in terrestrial ecosystems and the importance of 

natural emissions from undisturbed soils and vegetation is 

poorly understood. This situation persists because of 

experimental difficulties associated with ammonia 

measurements, the rapid gas-to-particle conversion of 

ammonia in the atmosphere, and the capacity of native soils 

and vegetation to act as both source and sink for atmospheric 

ammonia In the present paper, we attempt to summarize the 

current understanding of the natural sources and sinks for 

gaseous NH3 and the importance of natural emissions relative 

to anthropogenic emissions in the United States. We briefly 

review the physical and chemical processes that transform 

NH3 in the atmosphere, the major anthropogenic and 

potential natural sources of atmospheric NH3, and the 

techniques used to measure low concentrations and fluxes of 

atmospheric NH3. The available background concentrations 

I Also at Aeronomy Laboratory, NOAA, Boulder, 

Colorado. 

2 Also at Cooperative Institute for Research in 

Environmental Science, Boulder, Colorado. 

3Currently at Department of Biology, University of 

Stanford, Stanford, California. 

4Currently at NCAR, Boulder, Colorado. 

Copyright1992 

by the American Geophysical Union. 

Paper number 92GB02123. 

0886-6236/92/92GB-02123$10.00 

and flux measurements of ammonia in natural ecosystems are 

then described and used to infer upper limits for the emissions 

of NH3 into the atmosphere from these systems. While the 

magnitude of both anthropogenic and natural emissions of 

NH3 remain uncertain, it appears that unperturbed terrestrial 

ecosystems are generally more important as sinks rather than 

sources for atmospheric NH3. However, net emissions are 

likely from many eastern forests and other ecosystems 

exposed to large inputs of atmospheric sulfate. 

1. INlRODUCTION 

Ammonia (NH3) is the dominant gas-phase base in the 

atmosphere and the principal neutralizing agent for 

atmospheric acids. Ammonia thus influences the 

composition and pH of aerosols and cloud water and affects 

regional air quality and acid deposition patterns [Middleton, 

1987]. At low concentrations, atmospheric NH3 may be a 

valuable source of nitrogen for the vegetation in 

nitrogen-limited ecosystems [Hutchinson et al., 1972; Fahey 

et al., 1985] but is toxic to plants at high concentrations 

[Haynes, 1986; Vander Eerden, 1982] and emissions from 

anthropogenic sources have been implicated in the 

acidification of soils [Roelofs et al., 1987], forest decline 

[Van Breemen et al., 1982; NihlgArd, 1985; Schulze, 1989] 

and the eutrophication of surface waters [Hutchinson and 

Viets, 1969]. Consequently, there is considerable interest in 

the spatial and temporal distribution of ammonia 

concentrations and emissions. 

There is general agreement that livestock wastes and 

fertilizers are responSible for most of the atmospheric 

ammonia burden. Detailed emissions inventories have been 

developed for these and other anthropogenic sources in both 

Europe [ApSimon et al., 1987; BuijSinan et al., 1987] and 

North America [Harriss and Michaels, 1982; Heisler et al., 

1985; U.S. Environmental Protection Agency, 1989; Wagner 
460 

et al., 1986; Environmental Protection Service, 1985]. 

However, there remain large uncertainties about the role of 

undisturbed terrestrial ecosystems as a source for atmospheric 

ammonia and even the background concentrations in remote 

areas [Bowden, 1986]. There are three reasons for this. First, 

the plant-soil system can both emit or absorb gaseous NH3 

and a given soil or vegetation type may be a net sink in some 

areas and a net source in others. Second, scavenging of 

gaseous NH3 by acidic sulfate and nitrate aerosols 

complicates the interpretation of concentration and flux 

measurements in many areas [Harrison et al., 1989]. Finally, 

the net emissions of gaseous NH3 from natural sources and 

the concentrations in undisturbed ecosystems are generally 

small compared to anthropogenic sources and automated 

techniques sufficiently sensitive to make the continuous 

measurements needed to establish a large data base have not 

been readily available. 

The global distribution and behavior of atmospheric 

ammonia has been the subject of several reviews and budget 

analyses [e.g., Galbally and Roy, 1983; National Research 

Council, 1979; SOderlund and Svensson, 1976; Stedman and 

Shetter, 1983; Wameck, 1988; Schlesinger and Hartley, 

1992]. A concise summary and comparison of global budgets 

through 1983 is given by Wameck [1988] and a very recent 

global budget analysis published by Schlesinger and Hartley 

[1992]. The purpose of the present article is to assess the 

role of undisturbed terrestrial ecosystems as a source for 

atmospheric ammonia in the United States and the spatial and 

temporal variability of the background concentrations of NH3 

in the atmosphere. The techniques for measuring background 

concentrations and fluxes are reviewed, and the prospects for 

developing spatially and temporally resolved emission and 

concentration models discussed. 

2. SOURCES OF ATMOSPHERIC AMMONIA 

Ammonia is a ubiquitous product of biological activity. 

Sources in the plant-soil system include the mineralization of 

organic matter and reactions within the cells of living plants. 

Although most soil NJ4+ undergoes nitrification or is 

immobilized by vegetation, soil colloids, or clays, a small 

steady state pool is maintained in the soil solution. At 

sufficiently high pH, some of this ammonia exists as NH3 

which can be exchanged with the atmosphere. The soil 

solution will absorb gaseous NH3 when the atmospheric 

partial pressure exceeds the equilibrium vapor pressure and 

emit NH3 when the opposite is true. The NH3 vapor pressure 

(and hence the volatilization rate) is proportional to the 

dissolved NH3 concentration and increases exponentially with 

temperature according to Henry's law (cf. section 7.1 and 

Freney et al. [1983]). Thus the soil emissions rate and even 

the source or sink status can change with temperature and 

hence season or time of day. 

Inorganic fertilizers and animal excreta enhance soil 

emissions by directly increasing the NH3 or Nl4 + content of 

the soil solution [Hauck, 1983]. Loss rates from fertilized 

fields depend on the fertilizer type, soil, meteorological 

conditions, and the method of application [Fenn and Hossner, 

1985]. Surface application of urea ((NH2hCO) in animal 

urine or as fertilizer pellets results in the greatest losses since 

the subsequent hydrolysis 

Langford et al.: Ammonia Fluxes and Background Concentrations 

(I) 

also increases the pH of the soil within the urine patch or 

surrounding the fertilizer pellet. Schimel et al. [1986] found 

that ammonia N emissions average -=20% of urine N 

deposited in pasture grazing systems although this varies 

substantially with soil type, season, and pasture management 

practices. Much higher emission rates are likely when high 

densities of livestock are confmed in feedlots, barns, and 

poultry houses which are devoid of vegetation and therefore 

have little capacity for absorbing the generated ammonia. 

Ammonia is also released directly into the atmosphere 

through a variety of industrial activities. These include 

wastewater treatment, manufacturing processes, petroleum 

refining, coking, coal and oil combustion. Large amounts of 

ammonia are also volatilized during the burning of biomass 

associated with agricultural or forest management practices, 

paper manufacture, and wildfires. Several field and laboratory 

studies [Andreae, 1991; Griffith et al., 1989, 1991; Hegg et 

al., 1988, 1987; LeBel et al., 1988; Lobert et al., 1990, 

1991] suggest that pyrolysis of protein matter by smoldering 

fires produces large amounts of NH3 and CO, while flaming 

fires produce more oxidized species (NOx and C~). 

3. THE FATE OF AMMONIA IN THE ATMOSPHERE 

The interpretation of ammonia concentration and flux 

measurements is complicated by the interaction between the 

atmosphere-biosphere exchange process and transformations 

in the atmosphere. A simplified diagram illustrating the 

cycling of ammonia in the atmosphere and between the 

atmosphere and the biosphere is shown in Figure 1. Gaseous 

NH3 decreases with altitude above land surfaces and is larger 

over the continents than the oceans [Georgii and Gravenhorst, 

1977; Georgii and Miiller, 1974; Hoell et al., 1980; LeBel et 

al., 1985; Lenhard and Gravenhorst, 1980; Levine et al., 

1980; Tanner, 1982; Ziereis and Arnold, 1986]. Thus it 

seems likely that most of the ammonia present in the 

atmosphere above landmasses originates from terrestrial 

sources and that the oceans export little ammonia to the 

continents (see Quinn et al. [1987] for a review of the 

chemistry of ammonia in the remote marine troposphere). 

Gaseous NH3 does not absorb in the actinic ultraviolet and 

reacts only slowly with OH [DeMore et al., 1990]; the 

lifetime with respect to this process is -=3 months. This, 

coupled with the low concentrations of gaseous NH3 above 

the boundary layer, suggests that photochemical destruction is 

a relatively minor sink. However, it has been suggested that 

photooxidation of NH3 may be a potential source or sink for 

tropospheric NOx (i.e., NO+ N02) in remote regions 

[Logan, 1983; McConnell, 1973]. Photochemical box model 

calculations by the authors suggest that the importance of 

these processes depends heavily on the rate of reaction 

between NH2 and ~ and that they become entirely negligible 

within the boundary layer using the new upper limit 

determined in recent laboratory measurements by Tyndall et 

al. [1991]. 

The most likely fate for gaseous NH3 escaping the surface 

layer is incorporation into acidic aerosols and cloud water 

droplets in the boundary layer and lower troposphere 
Langford et al.: Ammonia Fluxes and Background Concentrations 

. . . . . . 

' ' ' ' ' 

461 

Wet Deposition 

F1g. 1. Schematic diagram illustrating the cycling of ammonia between the atmosphere and biosphere. Atmospheric ammonia 

may be present in the gas phase, in fine particulates (solid or liquid), and in liquid droplets (cloud or fog). The dominant processes 

are shown with solid lines; the dashed lines represent processes that are generally of secondary importance: evaporation ofNH3 

from ammonium sulfate, escape of gaseous NH3 the free troposphere, and adsorption of gaseous NH3 on plant surfaces. 

[Seinfeld, 1986]. The incorporation of ammonia into these 

bquid and/or solid phases displaces H+ as the primary cation 

and lowers the acidity thereby facilitating the uptake of 

weaker acids, like nitric acid, hydrochloric acid or organic 

ac1ds. The sequestering of ammonia by aerosols and cloud 

water occurs primarily through association with H2S04 to 

form ~HS04 and <NR!hS04. The equilibrium vapor 

pressure of ammonia above sulfuric acid is extremely small 

[Scott and Cattell, 1979; Tang, 1980] so that this process can 

be considered irreversible. Ammonia will be quantitatively 

Utrated from the gas phase leaving the particulate phase only 

partially neutralized if excess sulfate is present (i.e., 

([NH3]+~+])/ [S04=] < 2. Gaseous NH3 mixing ratios in 

excess of= 1 ppbv are likely only when there is sufficient 

ammonia to fully neutralize the particulate phase to 

<NR!hS04(i.e., ([NH3]+ ~+])/[S04=] ~ 2). 

Ammonia can also react with gas-phase HN~ to form 

~~. This association is reversible under atmospheric 

conditions [Stelson and Seinfeld, 1982] and ammonium 

mtrate can decompose releasing NH3. Because of the greater 

affmity of sulfate for ammonium, little nitrate will be 

Incorporated into the particulate phase until the sulfate is 

almost completely neutralized [Tang, 1980]. Thus NJ4N~ 

IS much more prevalent in the vicinity of western cities where 

the NO,/S~ emission ratio is large (e.g., Denver and Los 

Angeles) than in the eastern United States [Hidy et al., 1978; 

Wolff, 1984; Wolff et al., 1981]. In coastal areas, a similar 

reaction with hydrochloric acid forms ammonium chloride 

(NH4Cl) [Harrison et al., 1989]. 

The above processes lead to atmospheric lifetimes for 

gaseous NH3 of the order of several hours to a few days 

[Asman and Janssen, 1987; Georgii and Gravenhorst, 1977]. 

Long range transport is most important in the free 

troposphere above the planetary boundary layer. Over 

continental land masses, transport from the planetary 

boundary layer to the free troposphere occurs mainly by 

convective cloud pumping, frontal upwelling, orographic 

dynamics, and pumping associated with the diurnal variation 

in the height of the planetary boundary layer. The 

characteristic time for this transfer is of the order of 1 week or 

less [Luecken and Hales, 1986; Wameck, 1988] so that 

transfer out of the planetary boundary layer and long-range 

transport generally occurs after gas-to-particle conversion or 

cloud water scavenging. Release of gaseous NH3 following 

the evaporation of clouds is unlikely since most of the 

dissolved ammonia becomes associated with sulfate. Hence 

wet and dry deposition of particulate ~+ is the primary 

removal mechanism in most areas. 

4. TECHNIQUES TO MEASURE AMMONIA 

CONCENTRATIONS AND FLUXES 

4.1. Concentration Measurement Techniques 

Much of the uncertainty associated with atmospheric 

ammonia is due to the scarcity of reliable measurements. 

This in tum, is due to a lack of sensitive measurement 

techniques which are easy to use in the field and can operate 

automatically and continuously to establish large databases 

for statistical analysis. A major difficulty associated with 
462 

ammonia measurements is the presence of ammonia in 

gaseous (NH3), particulate (e.g., (NJ4)2S04 and NRtN~ ), 

and liquid (i.e., NJ40H in cloud and fog droplets) forms in 

the abnosphere. The partitioning between these three phases 

is highly variable and a complete picture generally requires 

simultaneous but separate measurements of at least the 

gaseous and particulate phases. A second major difficulty 

results from the tendency of ammonia to form strong 

hydrogen bonds with water and thus adsorb on the surfaces of 

most materials exposed to air. This can lead to high 

backgrounds and memory effects due to retention on sampling 

tubing and other plumbing. 

Most gaseous ammonia measurements reported in the 

literature 1rap ammonia with an acidic medium: acid bubblers 

[Breeding et al., 1973; Junge, 1957], acid-coated filters 

[Shendrikar and Lodge, 1975], or acid-coated denuder tubes 

[Ferm, 1979]. The solution from the bubblers or rinsed 

filters or tubes is then analyzed for ammonium via 

colorimetry, ion-specific electrode, or ion chromatography. 

The diffusion denuders developed by Ferm [1979] separate 

gaseous NH3 from particulate and droplet Nl4 + by exploiting 

the large difference in diffusion coefficient; gaseous NH3 in 

an airstream passing through the tube diffuses to and adheres 

to the acid-coated wall while the heavier particles pass straight 

through. Variants of this technique based on both cylindrical 

and annular designs have become the de facto standards for 

gaseous ammonia measurements. The detection limit depends 

on the sample time and denuder preparation, but levels below 

.. o.Ql ppbv appear to be possible for a 2-hour sample. 

Similar limits may be obtained with acid-coated filters which 

are also widely used, typically with a Teflon prefilter to 

collect the particulate phase. Filter packs are generally 

considered the standard technique for measuring particulate 

~+ and related species although more sophisticated 

techniques are under development and evaluation [Sickles et 

al., 1990]. 

Several intercomparisons of the diffusion denuder and filter 

pack techniques for measuring gaseous NH3 have been 

conducted in recent years. A field comparison of oxalic acid 

coated paper filters and diffusion denuders described by Gras 

[1984] indicated that volatilization of particulate ammonium 

collected on the Teflon prefilter led to high apparent gaseous 

NH3 in the filter pack technique. This result was also 

indicated in an intercomparison of filter pack, diffusion 

denuder, Fourier transform infrared (FTIR), and tungsten oxide 

denuder detectors in Claremont, California, during 1986 

[Appel et al., 1988; Wiebe et al., 1990]. In this case, 

ammonia concentration measured with the filter packs were 

approximately 1.5 times those measured with the oxalic acid 

denuders. Harrison and K.iitto [1990] also found evidence for 

particulate volatilization in the filter packs but concluded that 

the differences will typically be small compared to the natural 

variability of ambient ammonia These workers also stressed 

that the differences depend on the particular site, operator, and 

apparatus and thus do not reflect intrinsic limitations of the 

filter pack technique. 

Both filter packs and denuders offer the advantages of 

simplicity and relatively low cost, but the preparation and 

analysis of the sampler is labor intensive. Furthermore, great 

care must be taken to prevent contamination during storage 

and handling since ammonia is present in human breath and 

Langford et al.: Ammonia Fluxes and Background Concentrations 

perspiration as well as a variety of cleaning agents found in 

most laboratories. Thus the reproducibility and detection 

limit (and hence the necessary exposure time) obtained by 

these techniques largely depends on the skill and consistency 

of the operator. These restrictions make the acquisition of 

large data sets with sufficient temporal resolution to probe the 

parameters which influence abnospheric concentrations of 

ammonia difficult 

Several groups [Bos, 1980; Keuken et al., 1988] have 

developed automated diffusion denuder systems in which the 

acid solution coating is continuously replenished. The 

system developed at the Netherlands Energy Research 

Foundation (ECN) by Keuken et al. [1988] can operate for up 

to 8 hours without user intervention and detect other species 

as well as ammonia. Unfortunately, the detection limit for a 

30-min sample time is relatively high (~.3 ppbv). An 

alternate approach to automated detection uses "permanent" 

acid coatings and thermal desorption. McClenny and 

coworkers [Harward et al., 1982; McClenny and Bennet, 

1980] collected ammonia on Teflon beads or tungsten oxide 

tubes and detected the thermally desorbed product by 

photoacoustic spectroscopy with nominal a detection limit of 

=0.2 ppbv. Braman et al. [1982] developed a denuder tube 

coated with tungsten oxide/tungstic acid to collect ammonia. 

The collected ammonia is thermally desorbed and 

subsequently converted to NO on a hot gold surface. The NO 

is measured with a standard chemiluminescence detector. 

Field measurements based on this system are described by 

McOenny et al. [1982]. LeBel et al. [1985] used a similar 

instrument to measure venical profiles of ammonia above 

coastal Virginia and Maryland from an airplane, and Roberts 

et al. [ 1988] used this technique at Niwot Ridge, Colorado, 

and Point Arena, California. A similar instrument developed 

by Appel et al. [1984] and tested in the Oaremont 

inten:omparison described above agreed with the diffusion 

denuder measurements to within about a factor of 2 but 

exhibited a temperature dependent sensitivity. 

Thermal denuder systems based on other metal oxides have 

also been developed. Keuken et al. [1989] developed a system 

based on vanadium pentoxide while Langford et al. [ 1989] 

used molybdenum oxide. The former instrument employs an 

oxide coated annular quartz denuder while the latter uses the 

oxidized surface of a molybdenum rod suspended in a quartz 

tube. Both vanadium and molybdenum oxide directly oxidize 

the NH3 to NO or NOx upon desorption thus eliminating the 

need for a secondary convener. In ambient sampling tests, 

both of these instruments compared well with conventional 

denuders at mixing ratios above 1-2 ppbv. In general, metal 

oxide systems are reasonably simple to automate with good 

reproducibility and detection limits below 0.1 ppbv for 0.3-

to 0.5-hour sample times. The major drawbacks are related to 

the expense and long-term stability of the denuders in the 

field. These systems may also respond to aliphatic amines, 

although these species are generally present at very low levels 

relative to ammonia [Hutchinson et al., 1982]. 

Several other promising sampling techniques have been 

developed to measure gaseous ammonia. Abbas and Tanner 

[1981] described a fluorescence derivatization technique that 

collects ammonia in a continuously flowing acidic solution 

and monitors fluorescence from 1-alkylthiosoindole formed in 

the subsequent reaction with o-phthahaldehyde. The detection 
Langford et al.: Ammonia Fluxes and Background Concentrations 

limit is rather high (=0.3 ppbv for a 5-min response time); 

however, and the specifiCity toward gaseous NH3 has not 

been established. Rapsomanikis et al. [1988] used this 

analytical technique to quantify ammonia collected on filters 

or denuders. Farmer and Dawson [1982] used a cold plate to 

condense ammonia and other soluble gases in water, 

analyzing the condensate by colorimetry or ion-selective 

electrode. The extraction of atmospheric concentrations by 

this method is rather involved, and the estimated detection 

limit of 0.15 ppbv requires sampling times of several hours. 

Genfa et al. [1989] described an instrument in which 

ammonia diffuses through a porous membrane into water. 

The ammonium is then reacted with sodium sulfite and 

o-phthaldialdehyde and the resulting fluorescence from 

1-sulfonatoisoindole monitored. The technique offers great 

potential for general use since it is relatively inexpensive, 

appears to be highly specific toward ammonia, and has a 

detection limit of S0.05 ppbv for a 5-min sample time. At 

present, this technique has not been compared to other 

standard methods. Other instruments based on mist chamber 

techniques [Cofer et al., 1985] are under development in 

several laboratories and offer similar advantages. 

It is clear, however, that direct spectroscopic detection of 

gaseous NH3 without intervention of a collecting medium 

would be most desirable since this could provide both 

continuous measurements and unequivocal identification of 

the compound. Long path absorption techniques using C02 

differential absorption lidar (DIAL) [Force et al., 1985] and 

Fourier transform infrared spectroscopy [Biermann et al., 

1988] have been used but have very high detection limits (5 

and 1.5 ppbv, respectively). Tunable diode laser absorption 

spectroscopy (TDLAS), an in situ spectroscopic technique 

[MacKay and Schiff, 1984], has excellent sensitivity and 

specificity but the application at low concentrations is limited 

by ammonia losses on absorption cell walls. Recently, a 

photofragmentation/laser-induced fluorescence technique 

(PF/LIF) [Schendel et al., 1990] has been used which 

overcomes this problem by using very high flow rates 

(<::1000 L min-I) through the excitation cell. This technique 

appears to offer detection limits approaching 0.001 ppbv for a 

1-min integration time with no known interferences. 

However, in its present form, the system is large, expensive, 

and complicated. Nevertheless, the technique offers great 

potential as a transfer standard against which other 

measurements can be evaluated. 

The photofragmentation/laser-induced fluorescence 

(PF/LIF) technique was recently compared with citric acid 

coated diffusion denuders (CAD/IC), a filter pack/colorimetry 

system (FP/COL), a tungsten oxide denuder (DARE), and a 

molybdenum oxide denuder system (MOADS) in Boulder, 

Colorado, during 1989 [Williams et al., 1992]. This study 

indicated that the PF/LIF and CAD/IC systems agreed within 

the stated accuracies for all tests with no evidence of 

interferences. The FP/COL system was systematically lower 

due to temperature related losses on the Teflon prefilter. The 

DARE system agreed for separate inlet ambient sampling, but 

not for common inlet sampling due to retention of NH3 on 

the (Teflon) transfer line from the main sampling manifold 

The MOADS system showed a very strong correlation with 

the PF/LIF data for both common and independent sampling 

463 

configurations under clear conditions but had a systematic 

offset due to a calibration error. This system also exhibited a 

background level of =0.07 ppbv attributed to retention of 

NH3 on the Teflon inlets and responded to both gaseous NH3 

and free NJ4+ dissolved in cloud or fog droplets when the 

latter were present 

A main conclusion of the above study is that even short 

inlet lines retain ammonia and cause memory effects. In 

particular, Teflon appears to be unsuitable for ammonia 

sampling inlets. Thus sampling through any type of inlet 

should be done using high flow rates (i.e., <::100 L s-1) if low 

mixing ratios of ammonia ([NH3] :s; 1 ppbv) are to be 

measmed. 

4.2 Flux Measurement Techniques 

The only flux measurement techniques presently applicable 

to the measurement of ammonia fluxes from terrestrial 

ecosystems are those based on enclosure methods, gradient 

diffusion, or aircraft sampling. More sophisticated 

micrometeorological methods such as eddy correlation await 

the development of a sufficiently sensitive and fast ( ... 0.1 s 

integration time) NH3 detector. The techniques used to 

measure ammonia fluxes from agricultural systems have been 

reviewed by Harper [1990] and by Denmead [1983]. 

Enclosure methods are widely used to measure trace gas 

fluxes from soils and plants because of their relative 

simplicity. Fluxes are calculated from changes in the gaseous 

concentration within a small enclosure placed over the ground 

and purged with air containing a known concentration of the 

compound of interest. Similar techniques are widely used in 

laboratory studies of gas exchange. Since the enclosure 

technique measures fluxes from only a small area, replicate 

measurements spread over a larger plot are required to address 

the question of spatial variability in the field (at the other 

extreme, aircraft sampling interrogates such a large area that it 

is impossible to exclude nearby anthropogenic emissions). 

The chief criticism of the enclosure technique arises from the 

fact that the microenvironment within the box is not 

representative of the open field. For ammonia, this leads to 

serious consequences since soil fluxes are known to be 

sensitive not only to the atmospheric concentrations but also 

environmental quantities such as wind speed, temperature, 

evaporation rates, and dew formation [Freney et al., 1983]. 

Since ammonia is readily adsorbed on the materials used to 

make enclosure walls and sampling tubing (especially when 

condensation is present), the reliability of this technique at 

the low emission rates believed to be typical of unmodified 

field soils is highly questionable. For this reason, field 

measurements based on enclosure techniques are not included 

in the following discussion. 

Gradient diffusion techniques have been widely used to 

measure ammonia fluxes from agricultural systems 

[Denmead, 1983]. In this approach, ammonia fluxes are 

determined from simultaneous measurements of an ammonia 

concentration gradient 6[NH3]/6z and the heat or momentum 

flux. The latter is used to derive the eddy diffusivity (KtJ and 

the flux calculated as 

(2) 
464 Langford et al.: Ammonia Fluxes and Background Concentrations 

Pt. Arena 

Niwot Ridge 

WESTERN FORESTS 

~ Rocky Mountain 

1:·:·:-:·:·1 Pacific Coast 

~ Grasslands 

c::::::J Desert types 

Pawnee 

EASTERN FORESTS 

-Northern 

(::;:::::::::::::1 Central hardwood 

~Southern 

~Tropical 

Harvard Forest 

Thoreau's Bog 

Oak Ridge 

North Inlet 

Fig. 2. Distribution of ecosystem types in the contenninous United States. Background ammonia concentrations at sites 

representing most of these divisions are summarized in Table 1. Adapted from Brady [1974]. 

where a positive flux corresponds to emission from the 

surface. This approach does not perturb the plant-soil system 

or the influence of environmental factors on the exchange 

process. Gradient diffusion also horizontally averages the 

large spatial variability of ammonia fluxes over several 

hundred square meters. This characteristic is also a 

disadvantage in that it necessitates a large fetch and relatively 

simple terrain [Melillo, 1989]. As a result, most 

measurements have been made in grassland ecosystems. 1n 

some locations, the interpretation of gradient measurements is 

compounded by gas-to-particle conversion of ammonia in the 

atmosphere [Harrison et al., 1989]. Nonetheless, this 

technique appears to be the most reliable method available. 

5. MEASUREMENTS OF ATMOSPHERIC AMMONIA 

IN TERRESlRIAL ECOSYSTEMS 

Few NH3 concentration measurements have been made at 

sites that can be considered truly representative of undisturbed 

natural ecosystems. This is due not only to the difficulties in 

making NH3 measurements at low concentrations but also 

the interest in agricultural economy and air quality that 

motivates most atmospheric NH3 measurements. Similarly, 

an extensive literature exists for NH3 flux measurements in 

feedlots, cultivated soils, and fertilized and grazed pastures, 

while measurements of gaseous fluxes in unmodified 

grasslands and forests are virtually nonexistent In this 

section we summarize measurements made in reasonably 

undisturbed terrestrial ecosystems. 

5.1 Background ConcentraJion Measurements 

A compilation of summertime (May to October) 

measurements of gaseous NH3 at sites representative of the 

broad ecosystem types of the United States (cf. Figure 2) is 

given in Table I. The measurements are listed so as to form 

a rough longitudinal transect of the United States from west 

to east coasts. Only sites without permanent buildings and 

having relatively intact native vegetation are included. 

Several of the listed sites belong to the Long Term 

Ecological Research (L TER) network supported by the 

National Science Foundation (NSF). In comparing these 

results, it is important to note that the sampling techniques 

have not all been directly intercompared and that the 

measurements were made at heights ranging from 0.5 to 10m 

above the surface. Furthermore, the sample sizes vary widely 

and in some cases may be biased by nonrepresentative 

meteorological conditions. Finally, not all of these studies 
Langford et al.: Ammonia Fluxes and Background Concentrations 465 

TABLE I. Background Summer Ammonia Mixing Ratios in Natural Ecosystems 

Ecosystem Mean, Range, Sampling 

Site Type ppbv ppbv Technique Source 

Point Arena, Calif. Pacific coast 0.658 0.16-1.25 Tungsten oxide Roberts et al. [1988] 

denuder system 

Thurber Ranch, Ariz. desat 0.9 na condensation trap ion- Dawson and Farmer 

selective electrode [1984] 

Niwot Ridge, Colo. Rocky Mountain 0.42 0.11-1.66 molybdenum oxide Langford and Fehsenfeldc 

(0.62)b annular denuder 

Boulder, Colo. grassland 3.4 0.10-19.3 molybdenum oxide Langford and FehsenfeldC 

(4.4)b annular denuder 

Pawnee, Colo. grassland 11.4 2.5-47.0 molybdenum oxide Langford and Fehsenfeldc 

(11.8)b annular denuder 

Pawnee, Colo. grassland 12.2 6.8-17.6 acid trap colorimetry Zachariassend 

Konza, Kansas grassland 11.0 64.-16.9 acid trap colorimetry Zachariassend 

Kinterbish, Ala. southern forest 0.69 0.05-1.46 acid<Oated denudt7 Norton and Fehsenfeldc 

(0.42)b ion chromatography 

Oak Ridge, Tenn. central hardwood 0.30 0.11-1.88 molybdenum oxide Langford and Fehsenfeldc 

(0.30)b annular denuder 

Harvard Forest, Mass. northern forest 0.24 na acid<Oated denuder Tjepkema et al. [1981] 

ion-selective electrode 

Thoreau's Bog, Mass. freshwater wetland 0.33 na acid<Oated denuder Tjepkema et al. [1981] 

ion-selective electrode 

North Inlet, S. C. saltwater wetland 0.30 0.14-0.75 molybdenum oxide Langford and Fehsenfeldc 

(0.30)b annular denuder 

a The mean mixing ratio was = 0.32 ppbv when winds off the ocean are included. 

bMean mixing ratios during periods with T = 25 ± 2°C, horizontal wind speeds~ ms-1, and insolation ~350 W m-2. 

cPreviously unpublished results, 1992. 

dZachariassen, J., Ammonia exchange above natural grassland canopies, thesis, Colo. State Univ. Fort Collins, in preparation, 

1992. 

include nighttime measurements when the concentrations are 

typically lower (see below). The numbers shown in 

parentheses under the mean mixing ratios correspond to those 

periods with a temperature of 20t2°C, strong to moderate 

insolation, and little horizontal wind to advect NH3 from any 

nearby sources. 

The mean mixing ratios range from =0.3 ppbv at most of 

the coastal and forest sites to ~10 ppbv at two of the three 

grassland sites. The overall mean mixing ratio at Point 

Arena [Roberts et al., 1988], a coastal grassland =100 km 

north of San Francisco, decreases from =0.6 ppbv to =0.3 

ppbv when periods with westerly winds off the ocean are 

included. Dawson and Farmer [1984] measured gaseous NH3 

at a number of sites in the desert southwest and inferred the 

representative background concentration for this region to be 

0.9 ppbv. The wide range of mixing ratios observed at the 

Boulder grassland reflects a dependence on wind direction 

resulting from the location of this site at the boundary 

between the forested Rocky Mountains and the Great Plains. 

The value shown in parentheses is probably most 

representative. 

Gaseous NH3 typically exhibits both seasonal and diurnal 

trends. Figure 3a compares the average summer diurnal variation at Pawnee, Boulder, Niwot Ridge, Oak Ridge, and North 

Inlet and Point Arena. In each case, the NH3 mixing ratios 

have been normalized by the appropriate mean mixing ratio 

from Table 1. The NH3 mixing ratios are at a minimum in 

the early morning, peak near midday, and decrease during the 

night. The day/night ratio is =2 in all three cases. Figure 3b 

shows the seasonal dependence of the diurnal variation at 

Niwot Ridge in the Colorado mountains. The mean mixing 

ratio decreases from =0.4 to 0.2 ppbv between summer and 

winter. Even larger seasonal differences were observed by 

Tjepkema et al. [1981] at Harvard Forest in Massachusetts, 

and Cadle et al. [1985] observed very low wintertime values 

in northern Michigan. However, Dawson and Farmer 

observed only slight seasonal and diurnal trends in the desert 

southwest. 

The temporal variability of gaseous NH3 suggests a strong 

correlation with ambient temperature. This is illustrated in 

Figure 4 for Boulder, Niwot Ridge, Oak Ridge, and North 

Inlet. The mean mixing ratios differ by more than an order of 

magnitude, yet similar exponential temperature dependences 

are observed at all four of these very different .sites. As is 

discussed below, this observation provides a powerful 

argument for the control of gaseous NH3 in the lower part of 

the boundary layer by air-surface and/or gas-aerosol equilibria. 

The influence of gas-to-particle conversion on the form of 

ammonia in the atmosphere can be seen by comparing 

ambient measurements of gaseous NH3 and particulate Nl4 + 
466 Langford et al.: Ammonia Fluxes and Background Concentrations 

(a). 

1.75-.--------------------, 

1.5 

c 

0 1.25 

~ 

i 

8 

-:r:." z 0.75 Gl 

i 

;! 0.5 

0.25 ......... 

---

.....__ 

Pt. Arena, CA --+- oak Ridge, TN 

North Inlet, SC -a- Boulder, CO 

Niwot Ridge, CO -e- Pawnee, CO 

0~~~-r-~-T--~~--~--~~--T-~--~ 

0 4 8 12 16 20 24 

Local Standard Time 

(b). 

800,---------------------------------, 

700 

600 

c 

0 

= 500 ;: 

-. 

a; 

~ 400 

1:: 

• .e. 300 

-:r:." 

z 

200 

100 

Niwot Ridge, CO 

Winter 

0+-----.----~----------~---,----~ 

0 4 8 12 16 20 

Local Standard Time 

Fig. 3. (a) Diurnal variation of gaseous NH3 at several of 

the grassland, forest, and wetland sites listed in Table 1. The 

mixing ratios are normalized by the appropriate mean value 

listed in the table. (b) Seasonal dependence of the diurnal 

variation at Niwot Ridge, a forest site in the Colorado 

mountains. 

and S04=. Table 2 summarizes measurements from rural or 

suburban (but not necessarily "background") locations in 

California, Colorado, Alabama, Tennessee, North Carolina, 

Michigan, New York, Pennsylvania, and Connecticut made 

during both summer and winter. These studies represent all 

of the field campaigns at nonurban sites in the United States 

24 

to ~----------------------------------, 

J: 

z 

... 

0 N1wot Ridge, CO 

e Boulder,CO 

• N.lnle~ SC 

/j. Oak Ridge, TN 

to L---~~~~~~----~~~--~~~~ 

240 250 260 270 280 290 300 310 

T (K) 

Fig. 4. Temperature dependence of gaseous NH3 mixing 

ratios at four of the background sampling sites. The Boulder 

and Niwot Ridge plots include winter and fall measurements 

not shown in Table 1. Each symbol represents the average of 

five individual half-hour measurements spanning the 

temperature range indicated by the horizontal bars. The 

vertical bars represent the 95% confidence limit. 

that we are aware of with both gaseous and particulate 

measurements. Table 2 shows that while the total (gaseous + 

particulate) summertime ammonia concentrations are similar 

among most of the sites, the relative amounts of gaseous and 

particulate ammonia differ by more than two orders of 

magnitude. The gaseous and particulate phases of ammonia 

generally are uncorrelated. However, Figure 5 shows the 

rapid decrease in the gaseous fraction with increasing 

particulate sulfate. For sulfate levels below =-20 nmol m-3 

(=-2 IJ.g m-3), most of the ambient ammonia is in the gas 

phase during the summer. When sulfate levels exceed =200 

nmol m-3 (=20 IJ.g m-3), less than 5% of the ambient 

ammonia remains in the gas phase. The gaseous fraction 

decreases more rapidly with increasing [S04=] in winter due 

to reduced surface emissions of NH3 and the lower 

equilibrium vapor pressure of NH3 above the aerosol. 

5.2 Gaseous Flux Measurements 

A comprehensive tabulation of NH3 flux measurements 

from different terrestrial ecosystems is given by Schlesinger 

and Hartley [1992]. The measured emission rates range over 

more than 3 orders of magnitude with no obvious differences 

between forest, grassland, desert, and flooded ecosystems. 

While this result is seemingly at odds with the concentration 

measurements described above, many of these results are not 

based on direct flux measurements and reflect a wide range of 

sampling conditions with integration times ranging from 

minutes to weeks. In the following discussion we focus 

instead on a small number of gradient profile studies that 

shed light on the environmental and physical factors that 

influence the variability of gaseous NH3 fluxes. 
Langford et al.: Ammonia Fluxes and Background Concentrations 467 

TABLE 2. Regional Variability of Atmospheric Ammonia and Related Species in Rural and Suburban Locations 

Site Region NH3 NH4+ HNO:J NO:J· so4= Source 

Summer 

Newtown, Conn. New England 34 112 42 5 65 Keeler et al. [1991] 

Allegheny Mountain, Middle Atlantic SIS 172 78 9 200 Pierson et al. [1989] 

Pa. 

Oak Ridge, Tenn. Southeast 15 133 48 3 104 Lindberg et al. [1990] 

Oak Ridge, Tenn. Southeast 9 240 na 24 177 Langford, Norton, and Fehsenfeld• 

Kinterbish, Ala. Southeast 28 87 22 9 56 Norton, and Fehsenfeld& 

Boulder, Colo. Mountain 230 44 16 23 17 Langford, Norton, and Fehsenfeld• 

Niwot Ridge, Colo. Mountain 15 14 9 19 6 Langford, Norton, and Fehsenfeld• 

Winter 

State College, Pa. Middle Atlantic 2 141 40 21 60 Lewin et al. [1986]b 

State College, Pa. Middle Atlantic 3 65 18 2 26 Lewin et al. [1986]b 

State College, Pa. Middle Atlantic 11 73 22 4 36 Lewin et al. [1986]b 

Northern Mich. Great Lakes 2 43 11 19 28 Cadle et al. [1985] 

Boulder, Colo. Mountain 60 29 13 14 8 Langford, Norton, and Fehsenfeld• 

Niwot Ridge, Colo. Mountain 6 9 3 1 4 Langford, Norton, and Fehsenfeld• 

Lake Isabella, Calif. West 19 8 16 9 4 Jacob et al. [1986] 

Tehachapi, Calif. West 34 11 7 11 8 Jacob et al. [1986] 

Here 1 ppbv NH3 = 41 nmol m-3 = 0.7 Jlg m-3 at STP. 

•Previously unpublished results, 1992. 

bEntries refer to three measurement periods between January and March 1984. 

A landmark study in the field is that of Denmead et al. 

[1976], who measured NH3 fluxes above an ungrazed 

ryegrass/subclover pasture near Canberra in Australia during a 

2-day period in the spring. Ammonia was collected using 

acid traps at eight levels between 0.05 and 1. 71 m and 

exposed for 2 hours. The midday net emission rate was =0.8 

Jlg m-2 s-1 just above the soil surface and =0.06 Jlg m-2 s-1 

above the 0.7-m-high canopy. In contrast, an adjacent grazed 

pasture (50 sheep ha-l) released about 0.36 Jlg m-2 s-1 to the 

atmosphere during the same period. An earlier study 

[Denmead et al., 1974] at the latter site found fluxes of 0.84 

and 0.22 Jlg m-2 s-1 for fall daytime and nighttime 

measurement periods, respectively, or a daily mean of 0.30 

Jlg m-2 s·l. 

Harrison etal. [1989] determinedNH3 fluxes from a 

number of grass and crop-covered soils in England. Gradients 

were determined from filter pack measurements (2 to 4 hour 

exposures) at four levels between the surface and 3 m. 

Gaseous NH3 was collected on phosphoric acid impregnated 

paper filters, HN03 was collected on nylon filters, and 

particulate species were collected on Teflon filters. Great care 

was taken to account for the sometimes substantial deviations 

due to the reactions of NH3 with acidic species. The 

emission rates from grass-covered soils ranged from -0.03 to 

0.11 Jlg m-2 s-1 with no obvious difference between mown 

and unmown (5-1 0 em canopy height) surfaces. Diurnal 

patterns were not determined, but the measured fluxes exhibit 

a clear seasonal dependence with the largest emissions during 

the summer and fall and smaller positive or even negative 

.... 

:I: 

z 

+ 

fluxes in the winter (cf. Figure 6). The annualized daytime 

mean emission rate for grass-covered surfaces was =0.02 Jlg 

m-2 s-1 with a mean rate of =0.05 Jlg m-2 s-1 in the summer 

and fall. 

:~:"' 0.10 • :I: 

~ 

z • 

• 

0.01 +------,.....l..--...,------,---"'T'""------:j 

0 50 100 150 200 250 

so4= (nmol m"3) 

Fig. 5. Dependence of gaseous/particulate partitioning of 

ammonia on total atmospheric sulfate. Each point refers to 

one of the entries in Table 2. The open and closed symbols 

represent winter and summer values, respectively. 
468 

0.10 

0.08 ., 

E 

.;!, "' 

)( 

::J 

;;: 

~ 

:1: 

z 

c 

.. ., 

:; 

Month 

Fig. 6. Seasonal variation of ammonia fluxes determined by 

Harrison et al. [1989] at various grassland sites in England. 

Two recent studies by Langford and Fehsenfeld in 1988 and 

by Zachariassen and Schimel in 1989 examined the gaseous 

NH3 fluxes at a semiarid shortgrass steppe in the Pawnee 

National Grasslands of northeastern Colorado. This site is 

located on the U.S. Department of Agriculture (USDA) 

operated Central Plains Experimental Range L TER. The 

predominantly blue grama ground cover is quite sparse with a 

mean canopy height of only =0.15 m and frequent patches of 

exposed soil. The 1988 measurements were made using a 

molybdenum oxide annular denuder system at 0.5 and 2.0 m 

[Langford and Fehsenfeld, 1988]. A total of 881 half-hour 

averaged gradient measurements were made almost 

continuously over a 3-week period in June and July. Figure 

7a shows a time series for the daily averaged fluxes calculated 

from the measured gradients and micrometeorological 

measurements of heat and momentum flux provided by K. F. 

Zeller and W. Massman of the USDA Forest Service 

[Massman et al., 1990]. Net emission was observed on only 

7 out of 24 days, mostly during the first week of the study. 

A strong diurnal pattern was observed with mean emission 

rates of about 0.20 Jlg m-2 s-1 during the afternoons followed 

by deposition rates of approximately -0.10 Jlg m-2 s-1 during 

the nighttime and early morning hours (Figure 7b). The 

average (net) flux over the course of the study was -0.02 Jlg 

m-2 s-1 with daytime emissions almost exactly offset by 

nocturnal deposition. 

The 1989 study at Pawnee by Zachariassen and Schimel 

[1991] was designed to probe the role of the plant canopy on 

NH3 emissions. Measurements were made on eight days 

during four different parts of the growing season: green up 

(mid-June), peak biomass (late-June), dry down/senescence 

(early-July), and postsenescence (mid-to-late September). 

Gradients were determined from acid traps exposed for 2-hour 

periods at 1 and 3 m. The diurnally averaged mean flux rates 

observed during these periods were 0.004, -0.012, 0.007, and 

0.034 Jlg m-2 s-1, respectively. Patterns of nocturnal 

deposition and daytime emission consistent with those in 

Figure 7b were indicated for most of the sample periods. 

Langford et al.: Ammonia Fluxes and Background Concentrations 

Zachariassen and Schimel conducted a similar study during the 

summer of 1987 at the Konza Prairie Research Natural Area 

in eastern Kansas. The big bluestem, indiangrass, and 

switchgrass tallgrass prairie canopy at Konza is much denser 

than that at Pawnee. The NH3 concentrations were measured 

1 and 10 m above the soil surface. Samples were collected on 

8 days encompassing green-up, peak biomass, dry down and 

senescence, and postsenescence. The pattern was similar to 

that observed at Pawnee with diurnally averaged mean flux 

rates of 0.003, -0.017, 0.005, and 0.041 Jl& m-2 s-1, 

respectively, for the four periods. 

Several studies have examined the effects of grazing and 

fertilizer addition on the NH3 flux from grassland soils. A 

.. 

E 

"' 

0.2 

0.1 

.;; 0.0 

)( 

::J 

u:: 

... ::r:: ~.1 

z 

a; 

z 

(a). 

CftO'f""NM.IftCD,.._COQtO.-NC"I 

M---------.,....,....,....,....,....,....,....,....,....,...NNNN 

CD -~~~~~~~~~-------------­ ~~,.._,.._,...,...,.._,...,...,...,...,...,...,... 

Date 

(b). 

~~-------------------------------. 

.. 

E 

::r:: 

z 

E.. "' 

200 

100 

o l T 

+ r..l...U.. 'T' +o+++• 

-1oo ~.fTTTT 

-200 

0 1 2 1 6 20 24 

Local Standard Time 

Fig. 7. (a) Time series of diurnally averaged ammonia fluxes 

at the Pawnee Grasslands in northeastern Colorado during 

July of 1988. {b) Diurnal variation of the mean flux over 

the sampling period. The vertical bars represent the standard 

deviation of the mean. 
Langford et al.: Ammonia Fluxes and Background Concentrations 

comprehensive year long study of NH3 fluxes from a lightly 

grazed (4 beef cattle ha·l) tropical pasture in Queensland, 

Australia was described by Harper and coworkers [Harper et 

al., 1983a,b]. Profiles were determined from measurements at 

2 to 5 levels between 0.1 and 0.8 m above the soil surface. 

NH3 was collected using acid traps exposed for 2-hour 

intervals. The pasture was heavily fertilized four times per 

year with 374 kg N ha·l as urea, but the soil pH and~+] 

levels (and NH3 fluxes) returned to near background levels 

within =4 days after spring and fall applications. Diurnal 

cycling was then observed with much larger fluxes during the 

day than the night. Net deposition typically occurred during 

spring and summer nights, while small emissions (~0.02 ~g 

m·2 s·l) were sometimes observed during fall and winter 

nights. Short periods of deposition around sunrise and sunset 

were observed year-round but were most common during 

summer and fall. During the summer, NH3 flux was 

correlated most strongly with soil temperature, wind speed, 

and evaporation in that order. During the rest of the year, the 

correlation with evapotranspiration was strongest. Ranges for 

the net background fluxes were =-0.2 to 0.2, -0.02 to 0.6, 

0.06 to 0.8, and 0.2 to 0.3 ~g m·2 s·l during spring, 

summer, fall, and winter, respectively. 

The grassland results are consistent with measurements 

made by Harper et al. [1987] above a wheat crop grown in 

northeastern Georgia. Substantial NH3 emissions (=30% of 

the applied nitrogen) occurred immediately following the 

application of ammonium nitrate fertilizer, but decreased with 

time and the plant-soil system became a net sink for NH3 

(=-0.06 ~g m·2 s·l) within 30 days of application. The 

system became a net source again (=0.15 11g m·2 s·l) when 

the plants reached anthesis and continued through fmal 

senescence of the leaves, a total period of =40 days. More 

recent studies (L. A. Harper et al., Ammonia transport in a 

temperate grass system: Seasonal transport in relation to 

fertility and crop management, submitted to Agronomy 

Journal, 1993a; L. A. Harper et al., Nitrogen cycling in a 

clover system: Soil, plant, and atmospheric transport, 

submitted to Agronomy Journal, 1993b) show that net 

deposition also predominates in temperate ryegrass and clover 

fields once fertilizer inputs are immobilized. 

The gradient profile technique is much more difficult to 

apply in complex terrain [Fowler and Duyzer, 1989] and 

measurements from forest soils are virtually nonexistent. 

However, Baldocchi and Meyers [1991] recently described the 

special requirements for gradient profile measurements within 

an oak-hickory stand in the Walker Branch Watershed near 

Oak Ridge, Tennessee, and the molybdenum oxide annular 

denuder system and gradient profile method was used to 

measure the flux of NH3 there during September of 1988 (A. 

0. Langford et al., unpublished results, 1992). Walker 

Branch (mean elevation =300m) is administered by the 

Department of Energy and was one of the sites chosen for the 

United States/International Biological Program Eastern 

Deciduous Forest Biome Project supported by the NSF. 

Johnson and van Hook [1989] have reviewed the 

biogeochemical cycling at this site in detail. More than 400 

half-hour averaged flux measurements were made over a 3-

week period during the summer of 1988. The results of these 

flux measurements are shown in Figure 8. The analysis 

suggests a pattern similar to that at Pawnee but with very 

small average emission rates of~.001 ~g m·2 s·l during the 

"' 

E 

::t: 

z 

Cl 

.s 

)( 

:I 

u:: 

469 

.f 

ftt' f f ·1 

·2 

·3 

0 4 1 2 1 6 20 24 

Local Standard Time 

Fig. 8. Diurnal variation of the ammonia flux measured 

above the surface of an oak-hickory forest within the Walker 

Branch watershed near Oak Ridge, Tennessee, during October 

1988. The vertical bars represent the standard deviation of the 

mean. 

afternoons followed by deposition rates of =O.OOWg m·2 s-1 

during the nighttime and early morning hours. The average 

(net) flux over the course of the sampling period was 

indistinguishable from zero. 

6. MODELING AMMONIA EMISSIONS 

AND BACKGROUND CONCENTRATIONS 

IN TERRESTRIAL ECOSYSTEMS 

The flux studies described in section 5.2 suggest that most 

of the emissions of gaseous NH3 from terrestrial ecosystems 

are offserby local deposition. However, net emissions may 

occur under some conditions and the development of a simple 

model for predicting the spatial and temporal variability of 

NH3 emissions and background concentrations is of interest 

for regional air quality and acid deposition models. Several 

mechanistic models have been developed to estimate NH3 

emissions from various soil types [e.g .. Parton, 1981; 

Schimel et al., 1990; Fleisher et al., 1987]. Most of these 

models require detailed knowledge of soil parameters that are 

extremely variable and typically unknown for natural 

ecosystems and are thus not easily extrapolated to large areas. 

Dawson [1977] circumvented this limitation by relating soil 

parameters to more readily available climactic information and 

developed a model that predicts emissions from undisturbed 

soils comparable to those from anthropogenic sources. In 

this section we examine the soil emissions model developed 

by Dawson and the prospects for developing similar models 

to predict background concentrations ofNH3 in terrestrial 

ecosystems. 

6.1. Processes Controlling Ammonia Exchange 

Dawson and others have described the emission of NH3 

from soils in terms of a simple equilibrium model using a 

bulk transfer expression of the form 
470 

(3) 

Here p(NH3)5 is the equilibrium vapor pressure of NH3 in the 

soil pores and p(NH3)8 is the partial pressure of NH3 in the 

atmosphere. The exchange coefficient ksa describes the 

efficiency of transport of NH3 through the soil to the surface, 

through the laminar boundary layer above the soil, and away 

from the surface. A positive flux (i.e., F>O) corresponds to 

emission and occurs whenever the partial pressure in the soil 

exceeds that in the atmosphere. Similarly, the soil will 

absorb ammonia when the atmospheric partial pressure 

exceeds the equilibrium vapor pressure above the soil 

solution. 

The equilibrium vapor pressure of NH3 above the soil 

solution is related to the dissolved ammonia concentration by 

Henry's law 

(4) 

where [NH3]5 is the ammonia concentration in the solution, 

R is the ideal gas constant, and KH is the Henry's law 

coefficient. Since dissolved ammonia is partitioned according 

to the equilibrium 

(5) 

[NH3]5 will depend on both the total ammoniacal nitrogen 

and the pH of the solution. The NH3fNH4 + ratio changes 

markedly with pH and is virtually zero (S 0.001) in solutions 

with pH S 6 but approaches one (~ 0.90) in solutions with 

pH ~ 10. Both the equilibrium constant for (5) and the Henry 

coefficient vary with temperature [Dasgupta and Dong, 1986; 

Emerson et al., 1975]. Hence (5) can be expreSsed in terms 

of the dissolved ammonium concentration ~+ ] 5, pH, and 

temperature of the solution 

p(NH3)8 = 1.325 x 104 (~+] /[H+] ) exp(-10,378/D (6) 

Equation (6) shows that the equilibrium vapor pressure 

increases with the solution [N~+ ], pH, and T and predicts an 

approximate doubling of the NH3 vapor pressure for every 

5•c increase in temperature and a tenfold increase with every 

unit increase in pH. 

Qualitatively, (6) implies that soil emissions of ammonia 

should be greatest from neutral to alkaline soils in warm 

climates with moderate soil moisb.Jre and adequate porosity 

that have a steady supply of ammonium provided by rapid 

turnover of organic nitrogen (i.e., grasslands and savannahs). 

Emissions from the relatively acidic soils of forests should be 

negligible, despite the abundance of organic nitrogen, while 

emissions from the alkaline soils of desert regions will be 

limited by the lack of soil moisture and the availability of 

ammonium [Schlesinger and Peterjohn, 1991]. 

6.2. The Soil Emissions Model of Dawson 

The actual calculation ofNH3 emissions from natural soils 

is complicated by the spatial and temporal variability of the 

pH and ~+ ]5 of the soil solution and the complex nature 

of ammonia transport through the soil and subsequently into 

Langford et al.: Ammonia Fluxes and Background Concentrations 

the air [Milchunas et al., 1988]. The populations of both 

decomposing microflora and nitrifying bacteria, and, hence, 

ammonium levels, depend on the availability of organic 

nitrogen, soil moisture, and temperature (Schimel et al., 

1986; Haynes, 1986; Paul and Clark, 1989). Soil 

ammonium may also be taken up (and released) by growing 

plants, immobiliZed by organic matter, and chemically or 

physically adsorbed by minerals in clays. In very dry soils, 

ammonia may be trapped by sites which normally adsorb 

water. Transport of ammonia through the soil by gaseous 

diffusion or by convection and liquid diffusion of dissolved 

ammonia will be inhibited in waterlogged soils. Finally, 

significant volatilization requires a low ambient partial 

pressure of NH3 and sufficient wind to transport the NH3 

away from the surface to maintain a driving gradient 

Dawson [1977] empirically related soil ammonium 

concentrations to available biomass and the relative rates of 

decomposition and nitrification. The latter were derived as 

general functions of temperab.lre and soil moisture. 

Evapotranspiration was used as a surrogate for biomass and 

soil pH was correlated with annual rainfall. Finally, a soil 

water dependent gaseous diffusion coefficient was estimated to 

calculate the emission rates using (3) and the assumption that 

p(NH3)s >> p(NH3)8 • The model predicts substantial 

emissions from soils with a maximum at mid-latitudes, 

negligible polar(> 70° latitude) emissions, and lower tropical 

emissions. The average mid-latitude emission rate of .. 0.05 

JJ.g m-2 s-1 implies a net annual emission of about 4 Tg NH3 

yrl for the conterminous United States. 

Although, the measured midday grassland emission rates 

described in section 5.2 are similar to the average mid-latitude 

emission rate of ==0.05 JJ.g m·2 s·l calculated by Dawson, the 

model implies similar emission rates for forest soils. The 

model also neglects deposition and canopy uptake and thus 

predicts emission under all conditions in contrast to the 

observations. This follows from the assumption that the 

vapor pressure of NH3 in the soil always exceeds the 

atmospheric partial pressure. Dawson noted that the steady 

state expression used to estimate soil NH4+ was based on 

measurements from forest soils and underpredicted grassland 

~+ levels by about a factor of 20. In addition, he was well 

aware that the relationship between precipitation and soil pH 

ignores the influence of factors such as parent material and 

vegetation type and that a rainfall/pH relationship derived 

from measurements in California and India would not 

necessary reproduce the soil pH distribution of the United 

States. Dawson's model leads to ecosystem independent 

summer ~+]5 values of 3-5 mM, and mean soil pH values 

of7.0 and 7.1, respectively, for the 30°-40°N and 40°-50~ 

belts that contain most of the United States. Together with 

(6), these pH values and ~+]5 values imply soil NH3 

partial pressures of .. 250 ppbv at 20°C, well in excess of 

ambient concentrations thereby justifying the assumption that 

the soil is always a source ofNH3. However, Hanawalt 

[1969] observed large soil uptake for atmospheric 

concentrations as small as 40 ppbv. 

A quantitative test of these assumptions can be made using 

parameters determined from bulk soil samples (==10 cm3) collected by Williams and Fehsenfeld (1991) of the Aeronomy 

Laboratory at the grassland and forest sites listed in Table 3. 
Langford et al.: Ammonia Fluxes and Background Concentrations 471 

TABLE 3. Calculated Soil Vapor Pressures, Measured Ambient Partial Pressures, 

and Soil Emission Rates Based on Bulk Soil Samples (200C) 

lNl4tl. 

Site Ecosystem Type g/kg 

Pawnee, Colo. semiarid grassland 3.5 ± 2.0 

Boulder, Colo. temperate grassland I4.0 ± 7.0 

Niwot Ridge, Colo. montane coniferous I7.3 ± 7.I 

Kinterbish, Ala. temperate coniferous 5.6 ± 3.1 

Oak Ridge, Tenn. temperate deciduous I7.9 ± 8.5 

These samples were analyzed for pH (I :I paste), exchangeable 

NJ4+ (2 M KCl extract), and other parameters by the Soil 

Testing Laboratory of Colorado State University. These 

results are used along with field measurements of soil 

moisture and temperature to estimate the soil vapor pressure 

using (6). Exchangeable NRt+ (i.e., the ammonium 

remaining on the cation exchange complex (CEC) when the 

soil solution evaporates) was related to £NH4+ ls using the 

nonlinear Frewtdlich relationship of Cameron and Kowalenko 

[1976]. The calculated mean ammonium concentrations range 

from 1.2 to 2.4 mM or within a factor of 2 of Dawson's 

estimates. 

A much larger discrepancy arises between the estimated and 

measured soil pH. The values assumed by Dawson (pH =7) 

are not unreasonable for grassland soils which typically range 

from .. 5.5 to 8.5 but are much larger than the typical range of 

p(NHJ)s, p(NHJ) •• Flux, 

pH ppbv ppbv ng NH3 m-2s-l 

6.05 ± 0.44 II± I8 11.8 ± 4.1 0.03 ± 0.6 

5.90 ± 0.22 33 ± I5 4.4 ± 1.9 1.0 ± 0.5 

5.22 ± 0.38 4.4 ± 3.6 0.62 ± O.I9 O.IO ± 0.09 

5.28 ± 0.26 1.2 ± 0.7 0.42 ± 0.32 0.04 ± 0.05 

4.71 ± 0.26 1.7 ± 1.9 0.30 ± O.I5 0.02 ± 0.02 

forest soil pH (~.5 to 5.5). The measured pH values in the 

present study (4.7 to 6.I) are consistently lower than those 

predicted using Dawson's empirical relationship by .. 0.85 pH 

uniL A pH increase of 0.85 leads to a factor of 7 increase in 

NH3 vapor pressure while a pH increase of 2 corresponds to a 

factor of IOO increase. Thus Dawson's model underpredicts 

soil [H+] (and hence overpredicts the soil vapor pressure) for 

grasslands by about a factor of 7 and forests by a factor of 40 

to 400. The values of £NH4+ ]1 and pH determined from the 

soil samples give soil vapor pressures of •I and .. 35 ppbv at 

20"C for the forest and grassland sites, respectively, which are 

only slightly larger than the ambient mixing ratios (Figure 

9). Table 3 shows the mean fluxes calculated for those five 

sampling sites using the calculated soil equilibrium vapor 

pressures, the measured atmospheric partial pressures, and the 

gaseous diffusion limited exchange coefficient used by 

Dawson, are S0.001 JJ.g m-2 s-1. 

100.0,-------------------:-1 6.3. The Role of Vegetation in Ammonia Exchange 

>" .D 

Q. 

.s. 

... ::c 

z 

10.0 

1.0 

0 

• . I ; • I 

I • 0 

0 

• I : • 

. :. o I o 0 0 

. . . . : 

Pawnee 

0 0 0 ~t 0 0 I 1 I 

I o o 

0 0 t 0

I I -: 

• • Boulder 

o o ~ Niwot Ridge 

0 -~ + Klnterblsh 

oak Ridge 

0.1 +---,---,----..-,----..-.---.--.----..--; 

4.00 4.50 5.00 5.50 6.00 6.50 7.00 

Soil pH 

Fig. 9. Calculated soil vapor pressures (points) and mean 

ambient partial pressures of NH3 measured at Pawnee, 

Boulder, Niwot Ridge, Kinterbish, and Oak Ridge (large solid 

circles). The horizontal bars show the span of pH values at 

each site. The vertical bars show the standard deviation of the 

mean. 

The most serious limitation of Dawson's model is the 

omission of canopy effects. Vegetation is now known to 

play an important role in ammonia volatilization [Denmead et 

al., 1976; Lemon and Van Houtte, 1980]. Decaying 

vegetation is the primary source of the organic nitrogen 

mineralized to ammonium in native soils and the above 

ground biomass may lose NH3 directly to the atmosphere 

following proteolysis [Farquhar et al., 1983; Whitehead and 

Lockyer, I989]. Uptake of gaseous NH3 by growing plants 

is well documented [Aneja et al., 1986; Denmead et al., 1976; 

Hutchinson et al., I9n; Lockyer and Whitehead, I986; Van 

Hove et al., I987, 1989], and several studies [Denmead et al., 

I982; Farquhar et al., 1980; Lemon and Van Houtte, I980; 

Sinclair and Van Houtte, 1982] have shown that growing 

vegelation will emit NH3 when exposed to air with 

concentrations below some "compensation point" thought to 

be related to the partial pressure ofNH3 in the substomatal 

cavities. 

The processes controlling the concentration of NRt+ 

within the cells of growing plants is extremely complicated 

and the dependence of the compensation point on species and 

growing conditions is poorly understood. Typical values 

appear to be of the order of ~I ppbv for trees [Langford and 

Fehsenfeld, I992], .. 3-5 ppbv for herbaceous crop plants 

[Dabney and Bouidin, I990; Farquhar et al., I980; Lemon and 

Van Houtte, I980], ·and ~IO ppbv for wheat and perhaps other 
472 

grain filling grasses [Harper eta!., 1987; Lemon and Van 

Houtte, 1980; Morgan and Parton, 1989; Parton et al., 1988]. 

The compensation point has an exponential temperature 

dependence consistent with the presence of a steady state 

concentration of dissolved ammonia in the waterfilm present 

on the mesophyllic surfaces of the substomatal cavities 

[Farquhar et al., 1980]. The compensation point is 

hypothesized tO vary as a function of the balance between 

NJ-4+ releasing reactions (deamination, nitrate reduction, and 

senescence-induced proteolysis) and NJ-4+ consuming 

reactions (N transport, assimilation via glutamine 

synthetase). The balance shifts to the former with increasing 

plant maturity leading to higher intracellular NH3 

concentrations [Morgan and Parton, 1989; Parton eta!., 

1988]. 

The flux studies described in section 5.2 suggest that NH3 

exchange above grasslands is controlled primarily by canopy 

uptake and emission during the growing season, and that 

substantial net emissions are likely only during anthesis, 

flowering, and senescence of the canopy. Similar conclusions 

have been drawn from a number of studies in agricultural 

systems [Dabney and Bouldin, 1990; Lemon and Van Houtte, 

1980; Sinclair and Van Houtte, 1982]. Denmead et al. [1976] 

found that the emissions were much larger just above the soil 

surface(= 0.8 Jlg m-2 s·1) than above the 0.7 m high canopy 

(= 0.06 Jlg m-2 s·1) indicating that virtually all of the NH3 

emitted by the soil was reabsorbed by (or redeposited on) the 

canopy. This result is not unreasonable given the large 

surface area of a dense canopy compared to the underlying 

soil, the corresponding lack of ventilation near the soil 

surface, and the fact that the transport of gaseous NH3 

through even partially closed stomata is probably much faster 

than transport through the soil. 

6.4. Equilibrium Modeling of Background Concentralions 

The equilibrium nature of the exchange process implies not 

only that net NH3 emissions from terrestrial ecosystems will 

generally be small but also that the background 

concentrations of gaseous NH3 in many ecosystems should 

approach the characteristic temperature-dependent equilibrium 

vapor pressure of the soil or compensation point of the 

canopy. Thus it should be possible to use an approach 

similar to Dawson's to predict ambient concentrations. The 

general nature of this phenomenon is suggested by the similar 

temperature dependences of ambient NH3 at the different sites 

shown in Figure 4. This conclusion is also supported by 

Figure 9 which shows that the calculated vapor pressures at 

the Boulder and Pawnee grassland sites are very similar to the 

measured mean atmospheric partial pressures. The ambient 

mixing ratios at the forest sites are significantly lower than 

the soil vapor pressures, however, suggesting that the forest 

canopy compensation point is lower than the soil vapor 

pressure. The canopy shouid be particularly important as an 

NH3 regulator in forests and other ecosystems with dense 

vegetation. This hypothesis is supported by measurements in 

a montane/subalpine forest lying between Boulder and Niwot 

Ridge, Colorado [Langford and Fehsenfeld, 1992], which 

showed that the ambient concentration of NH3 near the top of 

the canopy approached the same temperature-dependent value 

when exposed to air containing NH3 concentrations differing 

Langford et al.: Ammonia Fluxes and Background Concentrations 

by more than a factor of 20. It seems likely that the temporal 

variability of ambient NH3 background concentrations can be 

represented by the temperature dependence for modeling 

purposes. However, development of a general model will 

require additional measurements to identify appropriate 

surrogates for the spatial variability. At present, we infer 

typical (to within a factor of two) background values of =10 

ppbv for grasslands, 0.5 ppbv for forests, and I ppbv for arid 

lands at 20'C. 

7. SIGNIFICANCE OF AMMONIA EMISSIONS 

FROM TERRESTRIAL ECOSYSTEMS IN THE 

UNITED STATES 

7.1. Magnitude of Natural Ammonia Emissions 

The measurements of NH3 concentrations and fluxes 

presented in the preceding sections suggest that the net 

emissions of NH3 from undisturbed terrestrial ecosystems are 

relatively small. Upper limits for these emissions from 

different ecosystems of the United States can be estimated 

using equation (2) and the background concentrations from 

Table I. Galbally [1975] used this approach to estimate the 

total flux ofNH3 to the atmosphere from land and ocean 

surfaces in the northern hemisphere assuming that the NH3 

concentration decreases by 50 to 90% between I m and 2 km, 

and long-term climatological average eddy diffusivities of 103 

cm2 s-1 at I m, lOS cm2 s-1 at 200m, and 2 x lOS cm2 s-1 

at 2 km. A total northern hemisphere flux of 130 Tg N yrl 

was obtained using constant surface mixing ratios of 7, 18, 

and 4 ppbv for temperate, tropical, and oceanic regions. In 

this section, we revise these calculations using more 

representative NH3 concentrations and estimate seasonal and 

diurnal effects. 

The results tabulated in Table I suggest mean daytime 

surface summertime mixing ratios of 1, 15, and 0.5 ppbv are 

representative of deserts and shrublands, grasslands, and 

forests, respectively. Nighttime mixing ratios are set at 50% 

of the daytime values, and wintertime mixing ratios at 50% 

of the corresponding summertime values for the calculations. 

It is assumed that gas-to-particle conversion and cloud 

scavenging reduce the mixing ratios to = 0 by I km and longterm average eddy diffusivities between the surface and 1 km 

are assumed to be 2 x lOS cm2 s-1 and 1 x 105 cm2 s-1 for 

daytime during the summer and winter, respectively (E.-Y. 

Hsie, private communication, 1991). Since stable nocturnal 

inversion layers are often formed, the mean nighttime eddy 

Jliffusivities are assumed to be =25% of the daytime values. 

This leads to estimated mean summer daytime emission rates 

of .. o.ot5, 0.2, and 0.007 Jlg NH3 m-2 s-1 for deserts, 

grasslands, and forests, respectively. The grassland rate 

compares favorably to the mean value actually observed at 

Pawnee. Wintertime emissions rates are 25% of the daytime 

rates, and nocturnal emission rates are .. 12% of the 

corresponding daytime rates. Since deserts, grasslands, and 

forests occupy 1.6, 0.95, and 2.3 x 106 km2 in the United 

States [Lamb et al., 1987; Olson eta!., 1980], this estimate 

yields annual emission rates of =0.2, 2.2, and 0.2 Tg NH3 

yr1 for deserts, grasslands, and forests, respectively, with two 

thirds of the emissions occurring during summer days. 

The above calculations suggest that grasslands are the only 
Langford et al.: Ammonia Fluxes and Background Concentrations 

natural ecosystems likely to be a significant source of 

atmospheric NH3 in the United States. However, the few 

available direct flux measurements suggest that much of the 

NH3 emitted by unfertilized and lightly grazed grasslands 

around midday during the warmer months is reabsorbed during 

the night Assuming a more ~nable net emission rate of 

0.2 Jlg m-2 s-1 for =3 hours per day over a growing season of 

100 days reduces the total emissions from U.S. grasslands to 

only 0.2 Tg NH3 yrl. The actual period with net emissions 

is probably even shorter spanning the period between anthesis 

(flowering), seed production, and fmal senescence and 

subsequent decomposition of the canopy. The Pawnee and 

Konza studies by Zachariassen [1991] imply net emissions 

of 0.67 and 0.74 kg ha-t during this period for the (temperate 

tallgrass) Konza and (semiarid shortgrass steppe) Pawnee 

grasslands, respectively. Extrapolation over the entire United 

States yields total grassland emissions of S0.07 Tg NH3 yr1. 

Undisturbed forest ecosystems should emit much less 

NH3. This situation is perturbed, however, by the addition of 

an atmospheric sink for gaseous NH3 in the form of 

unneutralized sulfate aerosols. As was pointed out by 

Galbally et al. [1982], the extremely low vapor pressure of 

gaseous NH3 above unneutralized sulfate aerosols depresses 

the ambient partial pressure below the compensation point 

leading to net emissions from the canopy. These processes 

will be particularly important in the eastern United States 

where high levels of unneutralized sulfate are typical (cf. 

Table 2). Under these conditions, the ambient partial pressure 

of gaseous NH3 above the canopy should reach a new 

"quasi-equilibrium" or steady state value intermediate between 

the compensation point and the equilibrium vapor pressure 

above the aerosol. Thus forests in the eastern United States 

are probably a net source for gaseous NH3. Similar 

arguments apply to grasslands exposed to unneutralized 

sulfate, but most grasslands are located west of the main 

sulfate source regions of the United States. 

473 

An upper limit for the emissions from forests can thus be 

calculated from the product of the canopy compensation point 

(y} and the stomatal conductance (g.). Assuming a 

representative value of g5 "' 0.17 mol m-2 s-1 for fully open 

stomata (equivalent to a deposition velocity of 0.4 em s-1 at 

sea level) and a mean compensation point of= 1 ppbv or =40 

nmol m-3, implies a volume emission rate of .. 0.15 nmol 

m-2 s-1 or =3 Jlg m-2 s-1 at 25 'C. Annual average forest 

emissions are calculated assuming mean flux rates of 3.5 and 

1.5 ng NH3 m-2 s-1 for summer and winter days, respectively, 

where the seasonal difference in emission rates reflects the 

temperature dependence of the compensation point. Winter 

emissions are probably much smaller, particularly in 

deciduous forests. Nighttime emissions are estimated to be 

=10% of the daytime rates since stomata typically do not 

close completely. Since forests occupy 2.34 x 106 km2 or 

31% of the total land area in the United States, these rates 

lead to an annualized forest emission of =0.1 Tg NH3 yr-1, 

similar in magnitude to the grassland emissions. 

7.2. Comparison to Anthropogenic Emissions 

Ammonia emissions from livestock wastes, fertilizer 

production and use, and other anthropogenic sources have 

been tabulated in detail for the United States [Harriss and 

Michaels, 1982; Heisler et al., 1985; U.S. EPA, 1989; 

Wagner et al., 1986] and Europe [ApSimon et al., 1987; 

Buijsman et al., 1987]. In addition, the Environmental 

Protection Service (EPS) has developed an inventory [EPS, 

1985] for Canada. The U.S. inventories are summarized in 

Table 4. Cattle and other livestock wastes account for 50 to 

80% of the total estimated anthropogenic ammonia emissions 

in all of these inventories. 

Total emissions in the published U.S. inventories range 

from 0.76 to 10.8 Tg NH3 yr1. The inventories designated 

as National Acid Precipitation Assessment Program 

TABLE 4. Comparison of Ammonia Emissions Inventories for the United States 

Emissions (Tg NH3 yr1) and Percentage of Total 

Livestock 

Combustion 

Industry 

Fertilizer application 

and land surfaces 

Wastewater 

Totalc 

1980 

NAPAP Inventory 

0.49 

0.23 

0.05• 

0.76 

(64%) 

(29%) 

(7%) 

1985 

NAPAP Inventory 

1.1 

0.07 

0.25 

0.05• 

(71%) 

(5%) 

(16%) 

(3%) 

0.07 (5%) 

1.5 

•Includes only the application of anhydrous ammonia fertilizer. 

bJncludes emissions from undisturbed land surfaces based on Dawson model. 

cValues may not sum to totals due to independent rounding. 

EPRI Inventory 

5.5 (50%) 

0.7 (6%) 

0.02 

4.61> (42%) 

10.8 

NASA Inventory 

2.1 (62%) 

0.7 (21%) 

0.2 (6%) 

0.4 (12%) 

3.4 

Sources: Placet et al. [1990], Wagner et al. [1986], U.S. EPA [1989], Heisler et al. [1988], and Harriss and Michaels [1982]. 
474 

(NAPAP) [U.S. EPA, 1989; Wagner et al., 1986] and 

Electric Power Research Institute (EPRI) [Heisler et al., 

1985] included detailed temporal and spatial resolution for the 

United States and partS of Canada. Most of the disparity 

results from large differences in animal emission factors and 

the included source categories. Thus the EPRI inventory is 

larger than the NAPAP inventories primarily because the 

EPRI emission factors for cattle are about 3 times higher than 

those used in the 1985 NAPAP inventory (40 kg animal-! 

yr-1 versus 12 kg animal-! yr-1) and 7 times higher than those 

used in the 1980 NAPAP inventory. The EPRI inventory 

also includes a large emission of ammonia from both 

croplands and unfertilized land surfaces such as pastures and 

rangeland. 

These disparities reflect the uncertainties inherent in 

extrapolating a small number of localized measurements to 

large scales and the dependence of NH3 volatilization on 

environmental effects such as temperature and humidity, as 

well as animal type, packing density, and surface. Although 

the emission factors used in the NAP AP inventory were ' 

chosen based on a careful evaluation of the earlier inventories 

and the reliability of the available data bases [Placet, 1990; 

Warn et al., 1990], the associated uncertainties lead to a low 

confidence rating from NAPAP. Despite the wide range of 

values, these inventories suggest that NH3 emissions from 

grasslands and forests in the United States will be relatively 

minor compared to the anthropogenic emissions associated 

with agricultural sources. 

7.3. Comparison to Emissions From Biomass Burning 

The emission inventories described above do not include 

NH3 production by biomass burning. Andreae [1991] 

estimates that biomass burning is responsible for "'12% of all 

ammonia emissions on a global basis with most of these 

emissions are associated with the clearing of tropical 

rainforests. In the United States, biomass burning is 

associated primarily with agriculture (e.g., annual burning of 

grasslands [Hobbs et al., 1991] or crop stubble), as a forest 

management procedure, or wildfires due to natural causes (i.e., 

spontaneous combustion or lightning initiated) or accidentally 

man-caused. According to Andreae [1991], "'80 and 250 Tg of 

dry matter are consumed in the United States and Canada each 

year through the burning of fuel wood and agricultural wastes, 

respectively. Wildfires probably consume an additional "'100 

Tg dm yr·1 in the United States [Logan, 1983; Yamate, 

1975]. The relative emissions ofNH3 and C~ have been 

examined for a variety of fires, and most studies have found 

NH /C~ emission ratios between "'0.1 and 3 mmol mol-l 

for both wetland and forest fu-es [Griffith et al., 1991] found a 

much larger value of "'8 mmol moJ·1 for burning sagebrush). 

Using Andreae's assumptions (cf. section 3) of a mean 

emission ratio of 1.3 mmol moJ-1, a carbon content of 45%, 

and that 88% of the carbon is volatilized as C~. we can 

estimate that ,5 x 10-4 gm of nitrogen is emitted as NH3 for 

each gram of dry matter burned. Alternatively, we can use the 

estimate of Lobert et al. [1991] that an average of 4% of the 

total N consumed in biomass burning is released as NH3 with 

a typical nitrogen content of 0.6% for the biomass burned in 

mid-latitude fu-es [Logan, 1983] to obtain a slightly smaller 

value of "'2.5 x 10-4 gm/gm. Thus we estimate an upper 

Langford et al.: Ammonia Fluxes and Background Concentrations 

limit of "'0.2 Tg NH3 yr-t for NH3 volatilization from 

biomass burning in the United States. 

7.4. Comparison to Ammonia Deposition 

Although it is not possible to separate the contributions of 

natural and anthropogenic sources, a lower bound for the total 

NH3 emissions for the continental United States can be 

estimated from the amount of~+ collected in wet 

deposition. Figure 10 shows the isopleths for ~+ 

concentrations in precipitation and the subsequent deposition 

as determined by the National Acid Deposition 

Program/National Trends Network (NADP/NTN) [1988] for 

1987. These data imply an annual wet deposition of 

approximately 1.5 Tg ~+ yr-1 for the conterminous United 

States. The spatial distribution reflects the presence of strong 

NH3 sources in heavily agricultural regions of the midwestern 

United States. The 1980 NAPAP inventory places 

anthropogenic emissions at only 50% of the measured wet 

deposition while the EPRI inventory (without soil emissions) 

is a factor of four larger. The 1985 NAPAP emissions are 

very similar to the wet deposition and the NASA emissions 

are a factor of 2 higher. Thus we can infer an absolute lower 

limit of"' 1.5 Tg NH3 yr-1 for NH3 emissions in the United 

States. 

Dry deposition inputs of gaseous and particulate ammonia 

are more difficult to estimate but are probably of similar 

magnitude. Deposition of excess gaseous NH3 to vegetation 

is very rapid with apparent deposition velocities of the order 

of 2.0 em s-1 [Duyzer et al., 1987]. This quantity is 

extremely variable, however, and includes contributions from 

both stomatal uptake and cuticular adsorption when the 

foliage is wet [Van Hove et al., 1989]. The latter is enhanced 

when unneutralized sulfate and nitrate are present on the 

surface [McLeod et al., 1990]. Hanson and Lindberg [1991] 

recently compiled a comprehensive review of dry deposition 

velocities for ammonia and other reactive nitrogen species and 

conclude that representative values for deposition of excess 

NH3 are 2.0 and 0.5 em s·l for forest and crop canopies, 

respectively. Dry deposition velocities for particulate~+ 

appear to be much smaller, ,0.1 to 0.2 em s-t. 

A recent modeling study by Asman and Janssen [1987] 

suggests that 65% of gaseous NH3 deposition occurs within 

5 km of agricultural sources and 90% within 30 km. A 

subsequent study by Asman and Jaarsfeld [1992] indicates that 

dry deposition of gaseous NH3 and particulate ~ + is 

responsible for 44 and 14%, respectively, of the total 

ammonia deposition in western Europe. Washout of NH3 

accounts for only 6% of the total since below cloud 

scavenging of gaseous NH3 is limited by the rate of transfer 

across the air-water interface and, as a consequence NH3 is 

generally not in equilibrium with raindrops [Asman and 

Janssen, 1987]; rainout of~+ incorporated into cloud water 

accounts for 36%. Harrison and Allen [1991] estimated that 

dry deposition of NH3 and ~ + accounts for 64 and 7%, of 

the total deposition in eastern England, with wet deposition 

responsible for the remaining 29%. These estimates are based 

on gaseous and particulate deposition velocities of 2.2 and 

0.15 em s·1, respectively, and become 45%, 11%, and 44%, 

respectively, if a gaseous deposition velocity of 1.0 em s-1 is 

assumed. 
Langford et al.: Ammonia Fluxes and Background Concentrations 475 

0_1 

0.2 

1987 annual precipitation-weighted mean ammonium ion concentrations (mg/L). 

Measured ammonium ion deposition (g/m2 ) for 1987. 

Fig. 10. Distribution of NJ4+ wet deposition collected by the National Acid Deposition Program/National Trends Network during 

1987. (a) Sample ammonium concentrations in milligrams per liter. (b) Corresponding deposition in g m·2. Note the 

preponderance of high ammonium levels in the central United States. 
476 

TABLE 5. Range of Estimated Ammonia Emissions 

for the Continental United States 

Source 

Livestock 

Fossil fuel combustion 

Industrial activities 

Fertilized cropland 

Biomass burning 

Native soils and vegetation 

Total 

Emissions are in Tg NH3 yr-1. 

Range 

1.1 to 5.5 

0.07 to 0.7 

0.02 to 0.25 

0.05 to 0.4 

s; 0.2 

s; 0.2 

1.2 to 7.25 

Limits on the total emissions of NH3 for the continental 

United States are presented in Table 5. The range of 

emissions are extracted from the estimates in Table 4 and the 

above discussions. These ranges were also determined, when 

appropriate, using the more recent NAPAP 1985 [U.S. EPA, 

1989] estimates in favor of the older NAPAP 1980 [Wagner 

et al., 1986] estimates. For each of the source categories, the 

range of emission is large, often a factor of 10. 

Unquestionably, however, the greatest emissions ofNH3 are 

associated with the raising of livestock. The range of NH3 

emission from livestock sources encompasses the NAPAP 

1985 [U.S. EPA, 1989] inventory estimate, 1.1 Tg NH3 

yr-1, and the EPRI [Heisler et al., 1985] inventory estimate, 

5.5 Tg NH3 yr-1. The sum of the sources provides a total 

range ofNH3 emission for the United States between 1.2 IRl 

7.3 Tg NH3 yr-1. The contribution from native soils and 

vegetation emissions is less than 20% of the total under the 

most favorable circumstances. 

8. NATURAL AMMONIA EMISSIONS AND ACID 

DEPOSmON 

Ammonia influences acid deposition both as a neutralizing 

agent for atmospheric aerosols and as a regulator of cloud 

water pH and hence S~ oxidation rates [Seinfeld, 1986; 

Behra et al., 1989]. Aircraft measurements by Tanner et al. 

[1984] show that the acidity of the atmospheric aerosol is 

generally much greater at altitudes of 1-2 km than at the 

surface. This results from the emission of S~ from tall 

stacks and the oxidation to sulfate occurring in cloud water, 

while ammonia sources are located on the ground. Thus 

precipitation is generally more acidic than the aerosol 

particles found near the surface. This can be seen from Figure 

11, which shows the relative amounts of NR.+ and so4= in 

both (1) precipitation and (2) particulates collected at the 

various rural sites listed in Table 2. In these cases, the 

particulate phase is more nearly neutralized (1s;[NR.+]/[S04=] 

s;2) than precipitation (fNH4+ ]/[S04=] ~.4). However, as 

was pointed out by Husar et al. [1989], the [NR.+] to [S04=] 

ratio in precipitation samples collected in the heavily 

agricultural upper midwest is typically »2. 

Langford et al.: Ammonia Fluxes and Background Concentrations 

(a). 

50 , 

, Precipitation , 

, 

, 

, 

, • Summer 40 , 1:1 0 Winter , 

, 

, 

, 

, 

, ~ 30 , 

3 , 

, 

, 

, 

, 

, • , l: 

z 20 , 

, 

, 

, 0 , • , 

, • 0 10 , • • ,. , 

, • 

20 40 6 0 80 100 

so: (J.!M) 

(b). 

500 

Particulates 

• Summer 400 0 Winter 

E 300 

0 

E 

.s 200 

• + ... 

l: 

z 100 

50 100 150 200 250 

so~ (nmol m· 3 ) 

Fig. 11. Relative concentrations of ammonium and sulfate in 

(a) precipitation and (b) aerosols at sites listed in Table 2. 

Figures 12 and 13 compare the spatial and seasonal 

dependences of the mean cation and anion concentrations in 

precipitation collected at 70 NADP/NTN sites in 1987. The 

data have been plotted in 5° longitude bins for clarity. Figure 

12a shows that acidic precipitation with H+ concentrations 

exceeding 10 J.LM (i.e., pH<5) is typical in the eastern United 

States, particularly during the summer months. This 

distribution reflects the preponderance of sulfur emissions in 

the east as is reflected in Figure 13a. The large 

summer/winter difference in the east results primarily from 

the much slower rate of S~ oxidation during the winter 

[Calvert et al., 1985]. In contrast, nitrate concentrations 

(Figure 13b) are less variable due to the greater dispersion of 

the sources and the much faster oxidation rates. Figure 12b 

shows that N~+ concentrations are generally greatest in 

precipitation collected in the midwest as would be expected 

given the large agricultural activity. However, there is very 

little seasonal variation in this region as might be expected 
:E 

.z, 

+ ... :I: 

z 

Langford et al.: Ammonia Auxes and Background Concentrations 

(a). 

on 0 on 0 on 

~ ~ '? .. .. 0 on 0 .;. .; .. .. .. .. .. Longitude 

(b). 

30 

20 

10 

0 0 on on on 0 on 0 ... .. 0 0 

~ ~ .. '? ':" 

";" ";" ";" 0 on .; on 0 

0 on 0 "' .. .. .. .. .. .. 0 0 "' 

Longitude 

Fig. 12. Concentrations of major cations, (a) hydrogen and 

(b) ammonium, in precipitation as a function of longitude and 

season. 

from the temperature dependence of ammonia emissions. 

This argues against large emissions from undisturbed 

grasslands and likely reflects the dominance of feedlot 

emissions since these often contain more animals during the 

winter than the summer months and high animal densities 

keep the surfaces clear of snow. In contrast, the NJ4+ 

concentrations in the east are not very different from those in 

the midwest during the summer months but are much lower 

during the winter. This could simply result from the seasonal 

variability of transport patterns but could also reflect the 

contribution of forest emissions. 

Although the total emissions of ammonia from forests is 

small on a continental basis, the emissions may be 

477 

significant in some parts of the east. Table 6 shows the 

forest emission rates calculated using (3) assuming p(NH3)a == 

0 and the forested area by state taken from [Olson et al., 

1980]. Also shown are the state-averaged anthropogenic 

ammonia emission rates estimated for domestic animal 

wastes, fossil fuels, and fertilizers by Harriss and Michaels 

[1982]. In most states, the calculated forest emissions are 

only ==10% of the anthropogenic emissions. In Ohio and 

Pennsylvania, the forest contribution is <5% of the 

anthropogenic emissions. By contrast, in New Hampshire 

and Maine the forest emissions account for 20-40% of the 

total emissions and may play a major role in the 

neutralization of atmospheric sulfate in this area. This 

contribution may be even greater if deposition of gaseous 

NH3 is faster than aerosol incorporation and most of the 

ammonia volatilized from localized anthropogenic sources 

such as feedlots and fertilized fields is absorbed by the 

surrounding vegetation. 

9. CONCLUDING REMARKS 

Ammonia remains one of the most poorly characterized 

species involved significantly in processes related to acid 

deposition. This situation can be attributed to the historical 

lack of suitable measurement techniques for gaseous 

ammonia, the rapid transformations and short lifetime of 

gaseous ammonia in the atmosphere, and the highly variable 

spatial and temporal nature of ammonia sources and sinks. 

As pointed out by Harriss and Michaels [1982], the primary 

reason for the poor state of knowledge is the lack of highquality data for either atmospheric concentrations or fluxes of 

ammonia. Recent developments in ammonia detection 

technology should improve this situation, but the new 

techniques are not yet perfected and are not universally 

available. Techniques to measure emission and deposition 

rates must also be developed further. However, it seems clear 

that wet or dry deposition and aerosol incorporation will often 

limit the lifetime of gaseous ammonia in the atmosphere to 

one day or less with long-range transport confmed to the 

particulate phase . 

Despite the limited number of measurements, it appears 

that the native plant-soil system will be a net sink or at most 

a relatively minor source of atmospheric ammonia compared 

to anthropogenic emissions. Net surface emissions from 

undisturbed soils and vegetation are likely to be negligible in 

most western United States forests and grasslands despite the 

high ambient ammonia levels found in the latter. In parts of 

the midwestern United States with strong localized sources of 

NH3 associated with animal feedlots, the soil and vegetation 

in surrounding areas will be net NH3 sinks. In the eastern 

United States and southeastern Canada, both forests and 

grasslands will be net sources of NH3 since unneutralized 

sulfate aerosols create a net upward gradient near the surface. 

However, the total emissions from these sources are probably 

~.2 Tg NH3 yr-1, or :5;20% of the total ammonia emissions. 

Nevertheless, these emissions may significantly impact 

sulfate neutralization in some localities and contribute to the 

general lack of equilibrium between atmospheric ammonia 

and sulfate aerosols. Detailed models of natural emissions 

will require additional measurements. 
478 Langford et al.: Ammonia Fluxes and Background Concentrations 

TABLE 6. Comparison of Anthropogenic and Natw-al Ammonia Emissions in the Eastern United States 

Anthropogenic 

State Domestic Animals• Fossil Fuels• 

Alabama 40.0 11.3 

Connecticut 3.4 1.4 

Delaware 0.9 1.2 

Georgia 42.3 19.4 

Kentucky 54.4 26.3 

Maine 4.8 0.7 

Maryland 8.5 9.1 

Massachusetts 2.9 3.5 

Michigan 32.2 30.2 

Mississippi 39.8 3.2 

New Hampshire 1.9 1.2 

New Jersey 3.1 6.2 

New York 42.8 18.7 

North Carolina 24.0 20.4 

Ohio 45.0 68.8 

Pennsylvania 42.7 63.3 

Rhode lsland 0.3 0.3 

South Carolina 12.0 8.8 

Tennessee 48.7 25.7 

Vennont 8.2 2.3 

Virginia 31.8 9.4 

West Virginia 9.1 30.4 

Emission rate in 1000 metric tons of NH3 per year. 

•Estimates from Harriss and Michaels [1982]. 

50 

40 

30 

:! 

.;!: 

II .. 20 0 

II) 

10 

0 0 

.. "' ~ 

"' 0 

~ ~ 

~ 

Fertilized Fields• 

3.2 

0.3 

0.7 

9.0 

13.5 

0.4 

2.4 

0.3 

10.6 

5.5 

0.1 

7.1 

2.3 

7.0 

19.0 

4.1 

0.1 

3.1 

3.4 

0.3 

3.1 

0.2 

(a). 

0 "' 0 

0 .. ':' .; "' "' .. .. .. Longitude 

Natw-al 

Forests Anthropogenic 

6.8 0.12 

0.6 0.12 

0.2 0.07 

8.0 0.11 

3.6 0.04 

5.1 0.86 

1.0 0.05 

1.0 0.15 

6.7 0.09 

6.6 0.14 

1.5 0.47 

0.8 0.05 

5.0 0.08 

6.5 0.13 

2.6 0.02 

4.9 0.04 

0.1 0.14 

4.3 0.18 

3.1 0.04 

1.3 0.12 

5.0 0.11 

3.4 0.09 

"' 0 "' ":' ":' ';" 0 "' 0 

.. .. .. 

Fig. 13. Concentrations of major anions, (a) sulfate and (b) nitrate in precipitation as a function of longitude and season. 
i" 

E. 

.. 0 

z 

Langford et al.: Ammonia Fluxes and Background Concentrations 

40 

30 

20 

10 

(b). 

• Winter 

.. 0 .. 

.. .. .. 0 .. 0 

.. .. .. 

Longitude 

Fig. 13. (continued) 

0 .. 

.. .. .. 0 

.. .. 

In conclusion, further improvements in our understanding 

of ammonia sources and sinks in the United States can be 

made through the following: 

1. Ammonia emissions from anthropogenic sources must 

be better characterized. Future inventories should be made in 

close collaboration with atmospheric and agricultural 

scientists to best evaluate animal emissions factors and 

fertilizer volatilization rates. New measurements in 

agricultural systems should be initiated to augment this 

process. 

2. Details of ammonia-cloud interactions, mass transport 

across aqueous droplets at low concentrations, and gas-toparticle conversion needs to be better understood and 

incorporated along with surface exchange into large-scale 

transport models for the United States. 

3. Future instrument development should concentrate on 

both fast response instruments for eddy correlation 

measurements and simple, inexpensive samplers that can be 

deployed at a variety of sites to better characterize the spatial 

distribution and variability of atmospheric ammonia. 

4. Field campaigns designed to characterize atmospheric 

ammonia should include supporting measurements of 

particulate ~+ and S04=, and in some cases gaseous HNO:J 

or HCI and particulate N03- or CJ-. Measurements of 

gaseous NH3 generally cannot be interpreted without 

knowledge of these quantities. 

Acknowledgments. A.O.L. and F.C.F. were supported by 

NOAA as part of the National Acid Precipitation Assessment 

Program and would like to thank Eric Williams for his soil 

analysis and assistance in the field measurements. D.S.S. and 

J.Z. were supported by NASA and NSF and wish to 

acknowledge the assistance of Bill Parton, Tom Kittel, Gary 

Peterson, Jack Morgan, and Tony Delany with their 

measurements and analyses. 

479 

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(Received March 10, 1992; 

revised August 28, 1992; 

accepted September 9, 1992.)