Ammonia Assessment from Agriculture: U.S. Status and Needs

Abstract

Recent studies suggest that human activities accelerate the production of reactive nitrogen on a global scale. Increased nitrogen emissions may lead to environmental impacts including photochemical air pollution, reduced visibility, changes in biodiversity, and stratospheric ozone depletion. In the last 50 yr, emissions of ammonia (NH3), which is the most abundant form of reduced reactive nitrogen in the atmosphere, have significantly increased as a result of intensive agricultural management and greater livestock production in many developed countries. These agricultural production practices are increasingly subject to governmental regulations intended to protect air resources. It is therefore important that an accurate and robust agricultural emission factors database exist to provide valid scientific support of these regulations. This paper highlights some of the recent work that was presented at the 2006 Workshop on Agricultural Air Quality in Washington, D.C. regarding NH3 emissions estimates and emission factors from agricultural sources in the U.S. and Europe. In addition, several best management practices are explored as the scientific community attempts to maximize the beneficial use of reactive nitrogen while simultaneously minimizing negative environmental impacts.

Full text

515
Recent studies suggest that human activities accelerate the
production of reactive nitrogen on a global scale. Increased
nitrogen emissions may lead to environmental impacts
including photochemical air pollution, reduced visibility,
changes in biodiversity, and stratospheric ozone depletion. In
the last 50 yr, emissions of ammonia (NH
3
), which is the most
abundant form of reduced reactive nitrogen in the atmosphere,
have signifi cantly increased as a result of intensive agricultural
management and greater livestock production in many
developed countries.  ese agricultural production practices
are increasingly subject to governmental regulations intended to
protect air resources. It is therefore important that an accurate
and robust agricultural emission factors database exist to
provide valid scientifi c support of these regulations.  is paper
highlights some of the recent work that was presented at the
2006 Workshop on Agricultural Air Quality in Washington,
D.C. regarding NH
3
emissions estimates and emission
factors from agricultural sources in the U.S. and Europe. In
addition, several best management practices are explored as the
scientifi c community attempts to maximize the benefi cial use
of reactive nitrogen while simultaneously minimizing negative
environmental impacts.
Ammonia Assessment from Agriculture: U.S. Status and Needs
Viney P. Aneja,* Jessica Blunden, and Kristen James North Carolina State University
William H. Schlesinger Institute of Ecosystem Studies
Raymond Knighton USDA
Wendell Gilliam and Greg Jennings North Carolina State University
Dev Niyogi and Shawn Cole Purdue University
B
 active, photochemically reactive, and radiatively
active nitrogen compounds in the atmosphere, hydrosphere,
and biosphere are collectively referred to as reactive nitrogen
(Galloway et al., 2003). Over the past few decades, human activities
leading to the production of reactive nitrogen from diatomic
nitrogen (N
2
) exceed that of nitrogen fi xation in the natural
terrestrial ecosystem at the global scale (Galloway et al., 2004).
Ammonia (NH
3
) is the most reduced form of reactive nitrogen.
It is also the most abundant alkaline constituent in the atmosphere
(Aneja et al., 2006a). Figure 1 illustrates the major processes (emis-
sions, chemical transformation, transport, and removal) that drive
the global cycle of NH
3
in the atmosphere (Aneja et al., 2006b,c).
In the past 50 yr, emissions and subsequent deposition of NH
3
have
increased signifi cantly in parallel with the development of intensive
agricultural management and increased livestock numbers (Sutton
et al., 1993). Globally, domestic animals are the largest source [32
× 10
12
g NH
3
–N (ammonia-nitrogen) yr
−1
] of atmospheric NH
3
,
comprising approximately 40% of natural and anthropogenic emis-
sions combined (Schlesinger and Hartley, 1992). Synthetic fertilizers
and agricultural crops together contribute 9 × 10
12
g NH
3
–N yr
−1
(12% of total emissions) (Schlesinger and Hartley, 1992).  e fi rst
Workshop on Agricultural Air Quality: State of the Science (Aneja
et al., 2006a) was structured to help scientists, industry, policymak-
ers, and regulators make optimal choices about issues confront-
ing agricultural practices to maximize the benefi ts and reduce the
detrimental environmental eff ects of current food, fi ber, and feed
production activities.  is paper focuses on issues surrounding am-
monia emissions, its transport, transformation, and fate.
Ammonia Emissions
Emissions of air pollutants, particularly ammonia, during agri-
cultural operations are an important emerging research area in the
U.S., best studied with interdisciplinary approaches that can inform
policymakers of the costs and benefi ts of various mitigation options.
Data on agricultural emissions of regulated pollutants, nuisance
odors, and fugitive dust often either do not exist or are insuffi cient
to develop appropriate policy nationally. Emissions estimates from
V.P. Aneja, J. Blunden, and K. James, Dep. of Marine, Earth, and Atmospheric Sciences,
North Carolina State Univ., Raleigh, NC, USA. W.H. Schlesinger, Inst. of Ecosystem
Studies, Millbrook, NY 12545. R. Knighton, USDA, Washington, DC, USA. W. Gilliam
and G. Jennings, Dep. of Soil Sciences, North Carolina State Univ., Raleigh, NC, USA. D.
Niyogi and S. Cole, Purdue Univ., West Lafayette, IN, USA.
Copyright © 2008 by the American Society of Agronomy, Crop Science
Society of America, and Soil Science Society of America. All rights
reserved. No part of this periodical may be reproduced or transmitted
in any form or by any means, electronic or mechanical, including pho-
tocopying, recording, or any information storage and retrieval system,
without permission in writing from the publisher.
Published in J. Environ. Qual. 37:515–520 (2008).
doi:10.2134/jeq2007.0002in
Received 1 Oct. 2007.
*Corresponding author (viney_aneja@ncsu.edu).
© ASA, CSSA, SSSA
677 S. Segoe Rd., Madison, WI 53711 USA
SPECIAL SUBMISSIONS

516 Journal of Environmental Quality • Volume 37 • March–April 2008
emerging agricultural producers in Southeast Asia (e.g., China
and India) and other developing countries are even more limited.
Table 1 shows the total 2002 estimated NH
3
emissions from vari-
ous animal husbandry operations in the United States (USEPA,
2005). Table 2 lists ammonia emission factors reported for the
Czech Republic, Denmark, and the Netherlands. Several other
European countries have also developed specifi c NH
3
emission
factors for intensive animal agriculture practices. Some countries
(e.g., the U.S.) are currently in the process of developing national
emission factors for both animal and crop agriculture (e.g., Table
3). Unfortunately, the process of developing appropriate emission
factors is hindered in the U.S., because relatively few experiments
with adequate spatial and temporal resolu-
tion have been conducted to estimate emis-
sions from animal production, especially
concerning dairy, beef cattle, and poultry.
Additionally, relevant studies often neglect
to report specifi c information (e.g., aver-
age animal weight, total area) necessary
to develop emission factors expressed as
[g NH–N kg
−1
d
−1
] in accordance with the
United States Department of Agriculture
(USDA) recommendations (Flesch et al.,
2007; Cassel et al., 2005; Gay et al., 2003).
Developing accurate emissions inven-
tories is also diffi cult because many factors
(e.g., seasonality, time of day, temperature,
humidity, wind speed, solar intensity, and
other weather conditions, ventilation rates,
housing type, manure properties or char-
acteristics, and animal species, stocking
density, and age) are involved in the genera-
tion and dispersion of ammonia. Emission factors therefore vary
both spatially and temporally, and all of the parameters discussed
above must be considered when used in air quality modeling. In-
formation on ammonia emissions from fertilizer applications and
confi ned animal feeding operations (CAFOs) provide only annual
averages and often disregard spatial and seasonal variations. Fur-
thermore, the uncertainties associated with these estimates are large
and emissions estimates applied for one set of conditions or for one
type of crop and animal feeding operation may not translate readily
to others. In general, agricultural emission factors from European
conditions have been applied for estimating ammonia emissions
in the U.S. To develop eff ective emission control strategies in the
U.S., it is imperative that appropriate U.S.-based emission factors
for both gases and particulate matter (PM) be developed.
Ammonia Transformation and Fate
Once released into the atmosphere, NH
3
has a relatively short
residence time of about 1 to 5 d (Warneck, 2000). When airborne,
it is either readily converted to aerosol or it is subjected to dry or
wet deposition processes. Ammonia is reactive with a variety of
acidic atmospheric species, including nitric acid (HNO
3
), hydro-
chloric acid (HCl), and sulfuric acid (H
2
SO
4
), which result in
the formation of ammonium aerosols, i.e., fi ne particulate matter
(aerodynamic diameter <2.5 μm). Due to the extended lifetime
of these aerosols (about 1–15 d), nitrogen may be transported to
previously pristine regions far from the pollutant sources (Aneja et
al., 2001). Assuming an atmospheric residence time of 6 d and a
wind velocity of 5 m s
−1
, Irwin and Williams (1988) estimate that
ammonium aerosols might travel as far as 2500 km; however, dry
and wet deposition may reduce this transport signifi cantly (Eris-
man and Draaijers, 1995).
Both ammonia and subsequently derived ammonium (NH
4
+
)
may be removed from the atmosphere through both wet and dry
deposition. Dry deposition occurs by diff usion of NH
3
in the
atmospheric boundary layer and the surface layer (Niyogi et al.,
Fig. 1. Atmospheric emissions, transport, transformation, and deposition of trace gases (Source:
Aneja et al., 2003, 2006b).
Table 1. Ammonia emission estimates in the United States for the
year 2002.†
Ammonia emissions
Source (tons/yr)
% of Total
emissions
Animal
Dairy(dairy cows and dairy heifers) 558,094 12.71
Beef(beef cattle, bulls, and calves) 656,648 14.95
Poultry(chickens and turkeys) 664,238 15.12
Swine(breeding and marketing pigs) 429,468 9.78
Sheep 24,835 0.57
Goats(milking and Angora goats) 14,028 0.32
Horses 71,285 1.62
Total animal 2,418,595 55.07
Fertilizer 1,140,396 25.97
TOTAL AGRICULTURE 3,558,991 81.03
Other
Chemical and applied product MFG 23,123 0.53
Fuel comb. elec. util. 30,256 0.69
Fuel comb. industrial 15,959 0.36
Miscellaneous 282,166 6.42
Fuel comb. other 17,602 0.40
Mobile sources 289,871 6.60
Waste disposal and recycling 25,770 0.59
Other industrial processes 148,288 3.38
Total other 833,035 18.97
TOTAL EMISSIONS 4,392,026 100.0
† Source: USEPA (2005).

Aneja et al.: Ammonia Assessment from Agriculture 517
2003). Wet deposition occurs by
below cloud scavenging (washout)
and rainout (in-cloud processes).
Fine particulate NH
4
+
is effi cient-
ly removed from the atmosphere
mainly through wet deposition.
Overall, wet deposition is more
important in regions with low
NH
3
emissions. Conversely, dry
deposition is more important in
regions of high NH
3
emissions
(Erisman and Draaijers, 1995;
Krupa, 2003, Niyogi et al., 2006).
Scientifi c information suggests
that reactive nitrogen (e.g., NH
3
)
is accumulating in the environ-
ment, and that nitrogen cycling
through biogeochemical pathways
has a variety of environmental
consequences including acidifi ca-
tion and eutrophication, photo-
chemical air pollution, reduced
visibility, ecosystem fertilization,
global warming, and stratospheric
ozone depletion (Galloway et al.,
2003). Lehmann et al. (2005) have
reported statistically signifi cant
increases in wet deposition of am-
monium and dissolved inorganic
nitrogen in much of the United
States over an eighteen year period
from 1985 to 2002. A number of
studies have evaluated the eff ects of
nitrogen deposition (Paerl, 1995,
1997; Paerl and Whitall, 1999;
Spiek et al., 1990; Sørensen et al.,
2003). Signifi cant excess nitrogen
deposition has occurred in the
eastern coastal areas of the United
States (Paerl, 1995). A particular
area of concern is the coastal rivers and their estuaries. Atmospheric
deposition of nitrogen compounds may contribute as much as 35
to 60% of total nitrogen loading to North Carolina coastal waters
(Paerl, 1995).  is excess nitrogen can result in toxic and non-toxic
phytoplankton blooms, which can lead to fi sh kills and reductions
of ‘clean water’ species (Paerl, 1995). In cooler climates of Europe
(e.g., Denmark) it is unlikely that atmospheric deposition events
trigger algae blooms directly (Hasager et al., 2003) within the
coastal waters surrounding Denmark. However, the deposition
amounts are high enough to increase chlorophyll production by
20% or more in these areas. Furthermore, the atmospheric deposi-
tion constitutes a large part of the overall load in these waters and is
therefore an important source for fi xed nitrogen (i.e., nitrogen tak-
en from its relatively inert molecular form (N
2
) in the atmosphere
and converted into nitrogen compounds such as NH
3
) (Spokes
et al., 2006). Soil acidifi cation is another problem experienced in
Northern and Central Europe, due to the high density of animal
operations. Van Breeman et al. (1982) identifi ed the deposition of
ammonium sulfate ((NH
4
)
2
SO
4
) as the main cause of soil acidifi ca-
tion in the Netherlands. Research conducted by Barthelmie and
Pryor (1998) in the Lower Fraser Valley, British Columbia, Canada
showed that NH
3
and NH
4
+
species and emissions play a particu-
larly critical role in visibility degradation. Fine particulate aerosols
have also been linked to human respiratory health problems. Stud-
ies suggest that the smaller the particle the greater the potential
health eff ect. For example, Lippmann (1998) found fi ne particles
(PM
2.5
) to be more toxic than coarse particles (PM
10
–PM
2.5
). Don-
aldson and MacNee (1998) examined ultra-fi ne particles (<100
nm) and found that toxicity increases as particle size decreases.
On the positive side, reactive nitrogen has contributed greatly to
the increase in food production, which becomes more important as
the human population is expected to grow to 8 to 12 billion by the
Table 2. Ammonia emission factors for individual animals categories for some countries in Europe
Animal category
Czech Republic 2002† Denmark 2005‡ The Netherlands 2006§
Emission factor Emission factor Emission factor
––––––kg NH
3
animal
−1
yr
−1
–––––– kg NH
3
animal place
−1
yr
−1
Dairy cows 27.9 26.92
Grazing 9.5
100% housed 11
Cows 16.2
Beef cattle 6–24 mo 7.2
Heifers 16.2
Heifers 6 mo-calving 5.49
Heifer calves < 6 mo 4.00
Calves 16.2 2.5
Bull calves 6–14 mo 8.45
Bull calves < 6 mo 3.51
Bulls 16.2 9.5
Other cattle 16.2 9.00 5.5
Farrowing sows (incl. piglets) 8.3
Dry and pregnant sows 4.2
Boars > 7 mo 5.5
Total Sows 17.44 7.60
Sucking-pigs 6.5 0.18
Large pen 0.75
Small pen 0.60
Pigs 8.3 1.01
Large pen 3.5
Small pen 2.5
Broiler breeders < 19 wk 0.250
Broiler breeders 0.580
Broilers 0.21 26.60 0.080
Layers 0.27 36.95
Layers < 18 wk (deep pit) 0.170
Layers + layer breeders(deep pit) 0.315
Turkey cocks and hens 0.92 29.00 0.680
Other poultry 0.21
Geese and dugs 0.73 7.54
Ducks; outside keeping 0.019
Ducks; inside keeping 0.210
Horses 8 8.32
Sheep 1.34 2.41
Goats 1.34 2.33
† Source: Zapletal and Chroust, 2006.
‡ Source: Gyldenkaerne, personal communication, 2007.
§ Source: Starmans and Van der Hoek, 2007.

518 Journal of Environmental Quality • Volume 37 • March–April 2008
end of the 21st century.  erefore, the challenge for the scientifi c
community is to fi nd ways to maximize benefi cial use of reactive
nitrogen while simultaneously minimizing adverse environmental
impacts. One way to approach this challenge is through the delib-
erate integration of reactive nitrogen research, management, and
control strategies. Integrated research and control strategies that
consider urban-rural air quality connections and interactions are
necessary for optimal ammonia/nitrogen management.
Best Management Practices
Assessments of best management practices (BMPs) that can
positively aff ect agricultural air quality and also provide economic
feasibility are still being examined. From the papers presented at
the Workshop and our review, a number of studies suggest the
best options to reduce ammonia emissions could be to reduce the
formation of ammonia and to increase nitrogen use effi ciency. An
example of a relatively simple solution to reduce ammonia produc-
tion is the reduction of crude protein in animal diets. Multiple
studies have found that reducing crude protein results in lower
NH
3
concentrations (Frank and Swensson, 2002; Powers et al.,
2006, 2007). However, crude protein is also an important com-
ponent of animal diets and it remains to be studied how it may
impact animal well being. In addition, when condensed tannin is
added to the drinking water of both cattle and sheep, less ammonia
is volatilized.  e amount of nitrogen in solid and liquid waste
was similar to that of regular water, but the nitrogen was nitrifi ed/
denitrifi ed rather than volatilized into ammonia
(Kronberg, 2006). However, if this is not possible
in current facilities, several BMPs and/or emission
reduction options to curtail ammonia emissions from
agricultural sources have been tested, mainly dealing
with emissions from cattle and swine. A relatively
simple solution was undertaken by Lefcourt and
Meisinger (2001), who tested the addition of alum
and zeolite to cattle slurry in an eff ort to curb the
volatilization of ammonia. When alum was added
at 2.5 and 6.25%, there were reductions of 58 ± 6%
and 57 ± 10%, respectively. Slightly lower reductions
were seen with the addition of zeolite, with additions
of 2.5 and 6.25% resulting in reductions of 22 ± 6
and 47 ± 10%. Similar tests were performed by Berg
(2006), with attempts to lower ammonia emissions
from cattle slurry by acidifying it with lactic and
nitric acid. Lactic acid was applied to reach pH levels
of 5.73, 5.14, and 4.18, yielding decreased emissions
of 65, 72, and 88%, respectively. Nitric acid reduced
emissions by about half of the success seen with lactic
acid. When nitric acid was added to reach pH levels
of 5.20 and 4.49, emissions were lowered by 29
and 49%, respectively. A side eff ect is that although
ammonia emissions are reduced, the total reactive
nitrogen production increases.  e costs of the addi-
tive (such tannin, alum, and zeolite) would depend
on the scale of operation and is not considered in the
reported studies.
More approaches to reduce dairy cattle emis-
sions include changes in fl oor design and ventilation. It was
found that if a fl at fl oor in a cattle barn was scraped 96 times
per day there was a 5% reduction in the ammonia emitted
from the barn. Yet, with a sloped fl oor, reductions were much
greater. With 12 scrapings per day, there was a 21% reduction
and, with 96 scrapings per day, there was a 26% reduction
(Bramm et al., 1997).  e study does not report on the im-
pact on other waste-handling emissions (e.g., lagoons).
Using a fi ltered, custom built, double-polytube ventilation
system in their calf barns, Hillman et al. (1992) were able to
reduce the ammonia concentration from about 5.7 mg m
−3
in
the air to about 2.9 mg m
−3
and below; however, the impact
on the total emissions remains to be studied.
Both simple and more detailed BMPs have been tested for the
reduction of ammonia from swine sources. A simple solution was
the addition of the manure additive Alliance to swine manure,
resulting in a 24% reduction in ammonia emissions (Heber et al.,
2000), while an example of a more complicated solution involves
ventilation and indoor air climate control (Hartung, 2006).
 ere was a reduction of 10 to 14% of ammonia when there was
a reduction of indoor air temperature and ventilation rate. Yet,
to obtain these results an evaporative indoor air cooling system
with an “optimization of the fogging control with regard to a
continuously complete evaporation of water” was needed. Other
solutions are more complicated in design and setup, but are easier
to use. Using biotrickling fi lters for the manure, Hansen and
Table 3. Ammonia emission factors from animal and crop agriculture in the US
Animal agriculture emission factor (kg-NH
3
animal
−1
yr
−1
)
Animal
USEPA
NEI 2002† Battye et al., 2003‡
USEPA
Battye et al. 1994§
Dairy cow 38.1 28 40
Beef cow 10.2 27 (steers)
On feed 20.8
Grazing 4.3
Pigs 6.5
Sow 16.4 16
Finishing pig 6.4 7
Poultry .27
Laying hen 0.31 0.31
Broiler 0.28 0.17
Sheep 3.4 1.34 3.4
Horses 12 8.0 12
Crop agriculture emission factor (kg NH
3
/Mg N)
Fertilizer Aneja et al. 2003¶ USEPA; Battye et al. 1994§
N-P-K 48 48
Nitrogen solutions 30 30
Ammonium phosphates 48 48
Anhydrous NH
3
12 12
Urea 182 182
Ammonium nitrate 25 25
Other straight nitrogen 30 30
Ammonium sulfate 97 97
Aqua NH
3
12 12
Ammonium thiosulfate 30 30
† Source: USEPA, 2004.
‡ Source: Battye et al., 2003.
§ Source: Battye et al., 1994.
¶ Source: Aneja et al., 2003.

Aneja et al.: Ammonia Assessment from Agriculture 519
Jensen (2006) were able to reduce both odor and ammonia emit-
ted from manure. Loyon et al. (2006) found that with the use of
a storage spreading system with biological treatment of manure,
there was a 30 to 50% reduction with separated manure and a
68% reduction with unseparated manure.
Emission controls of ammonia down to the farm level and
the optimization of agricultural production methods have
been evaluated in Denmark using modeling frameworks (Am-
belas Skjøth et al., 2006).  ese emission models are being
used for explaining variations and trends in ambient ammonia
concentrations as a result of changing agricultural practices
(BMPs) brought about by national regulations and policies.
Promising results have been reported for reducing am-
monia from swine manure through the use of an “engineered
system,” i.e., a treatment plant with solid-liquid separation.
Szögi and Vanotti (2006) reported a 73% reduction in am-
monia emissions from the implementation of such a system.
Vanotti (2006) found that when manure from such a system
was applied, there was a 98.8% reduction in greenhouse gas
emissions, as well as a potential for additional income of $9,100
to $27,500/yr (approximately $0.91/fi nished pig) from imple-
menting cleaner technology through the Supersoil program. In
addition, when organic fertilizers with gypsum are applied, they
can reduce ammonia volatilization by 11% (Model, 2006).
Conclusions
In the U.S., there are currently no regulations or incentive
programs that require reductions in ammonia emissions. It is
striking when comparing the diff erent pollutants, sulfur dioxide
(SO
2
), oxides of nitrogen (NO
x
), volatile organic compounds
(VOCs), and ammonia (NH
3
), that the extensive control mea-
sures applied to SO
2
, NO
x
, and anthropogenic VOCs have not
been extended to ammonia, which has similar or higher contri-
butions to e.g., PM exposure, visibility, or acidifi cation/eutrophi-
cation. However, future policies and control measures are neces-
sary to successfully decrease ammonia emissions and its related
problems. Current research problems related to ammonia include
the quantifi cation of agricultural point and nonpoint sources and
its temporal variation; re-emission in high emission/deposition
areas; the atmosphere-biosphere exchange of ammonia and its ef-
fect on SO
2
deposition; the quantifi cation of landscape processes;
low-level dispersion processes; the primary and secondary emis-
sions of PM; and the gas-to-particle conversion to PM fi ne.
In the past, policies to control criteria pollutants mainly focused
on single pollutants, while addressing, in general, single eff ects.
Today, multi-pollutant/multi-eff ect approaches are being consid-
ered e.g., in Europe, the Gothenburg Protocol under the United
Nations Economic Commission for Europe (UNECE) Conven-
tion on Long-Range Transboundary Air Pollution, and the EU
National Emission Ceiling Directive, thus off ering unique op-
portunities for the development of ammonia abatement measures
using integrated approach strategies across multiple media.
Generally, limited data exist for estimating agricultural emis-
sions of air pollutants (e.g., NH
3
, H
2
S) and public nuisances (e.g.,
odors, fugitive dust). Credible estimates of air emissions from
CAFOs are also complicated by factors that aff ect the amounts and
dispersion of emissions in the atmosphere. Emissions that occur
throughout the food production system must be quantifi ed accu-
rately to address critical air quality issues (Aneja et al., 2006c).
 e papers that are part of this special section address emis-
sions and fate of both gases (e.g., greenhouse gases, VOCs, etc.)
and particulate matter. New technologies for measurements and
analysis of trace gases, and transport and transformation model-
ing of these pollutants are explored.  e BMPs being studied for
gas and odor reductions and agricultural air quality, in general,
appear to be in their infancy and more comprehensive studies are
needed that can develop large scale, economically viable practices.
Acknowledgments
We acknowledge support from the Cooperative State
Research, Education, and Extension Service (CSREES), USDA
National Research Initiative Competitive Grants Program,
contract 2005-35112-15377. We thank Ms. Sally Shaver, U.S.
Environmental Protection Agency; Mr. Bill Battye of ECR,
Dr. Jan Willem Erisman, the Netherlands, and Dr. Ole Hertel,
Denmark for their constructive comments.  is paper was
presented at the International Conference on Ammonia in
Agriculture: Policy, Science, Control and Implementation, Ede,
 e Netherlands, 19–21 Mar. 2007.
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