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In: Environmental Chemistry of Animal Manure

Editor: Zhongqi He

ISBN 978-1-61209-222-5

© 2011 Nova Science Publishers, Inc.

Chapter 6



Pius M. Ndegwa1,*, Alexander N. Hristov2 and Jactone A. Ogejo3


Ammonia (NH3) volatilization is one of the most important pathways through which

nitrogen (N) is lost from animal manures. Ammonia volatilization is a critical issue because

its loss not only reduces the fertilizer-value of the manure but, in most cases, also has

negative impacts on the environment. Agriculture is believed to be the largest source of

global NH3 emission, with the majority of emissions (~80%) estimated to originate from

animal manures. Potential adverse consequences associated with NH3 emission to the

environment include: respiratory diseases caused by exposure to high concentrations of

secondary fine particulate aerosols formed from NH3 (commonly referred to as PM2.5); nitrate

contamination of drinking water; eutrophication of surface water bodies manifested in

harmful algal blooms and decreased water quality; vegetation and ecosystem changes caused

by excess N deposition; and soil acidification through nitrification and leaching. Reducing

NH3 emissions from animal manures can ultimately mitigate these environmental impacts. In

order to design effective techniques for reducing NH3 loss from animal manures, it is

important to first understand the fundamental mechanisms involved in such emissions. This

Chapter is divided into two sections. Section 6.2 presents processes and mechanisms leading

to NH3 production and volatilization from animal manures. Section 6.3, on the other hand,

summarizes basic principles of the techniques and case-studies, either already developed or in

development, for minimizing NH3 emissions from animal manures.


Corresponding author: ndegwa@wsu.edu; Phone: 509.335.8167; Fax: 509.335.2722.

Biological Systems Engineering, Washington State University PO Box 646120, Pullman, WA 99164, USA


Dairy and Animal Science Department, Pennsylvania State University, University Park, PA 16802, USA


Biological Systems Engineering, Virginia Tech. 212 Seitz Hall, Blacksburg, VA 24061, USA



Pius M. Ndegwa, Alexander N. Hristov and Jactone A. Ogejo


6.2.1. Processes Responsible for Ammonia in Manure

In monogastric animals; dietary N concentration is the primary factor determining N

excretions and NH3 volatilization losses from manure. The efficiency of utilization of dietary

N in pigs, for example, ranges from 30 to 50% (Jongbloed and Lenis, 1992; Arogo et al.,

2001), which is comparable to the efficiency of utilization of dietary N in the ruminants (see

following paragraphs). In general, overfeeding of protein and amino acids imbalance are the

main reasons for this low efficiency. Numerous studies have demonstrated reduced NH3

emissions from pig manure with reduced dietary crude protein (CP) intake (Le at al., 2009).

Meticulous balance of animal needs and amino acids supply can dramatically improve N

utilization efficiency and consequently reduce NH3 emissions from manure. For example,

feeding low protein diets supplemented with individual amino acids to meet, but not exceed,

animal requirements has produced remarkable improvements in the efficiency of dietary N

utilization for production purposes in pigs (Baker, 1996).

Nitrogen metabolism in ruminants is a more complex process than in monogastric

animals because of extensive breakdown and modification of proteins in the reticulo-rumen.

The ruminant animal is unique in its ability to convert feed N into microbial protein. The

metabolizable protein needs of the ruminant are met primarily from two sources of amino

acids: microbial protein synthesized in the rumen, and un-degraded feed protein in the rumen;

with a small contribution from endogenous protein secretions. Microbial protein amino acid

composition is very similar to the amino acid composition of tissue and milk protein (NRC,

2001; Lapierre et al., 2006), which makes it an ideal source of amino acids for the animal.

Feed protein that by-pass ruminal degradation, however, may not provide digestible essential

amino acids in the quantity or ratios sufficient for maintenance or production needs. Even if

the amino acids absorbed in the gut match closely the amino acid requirements, the liver

significantly modifies the amino acid profile of metabolizable protein available to the animal

(Blouin et al., 2002). During this metabolism, amino acids are deaminated and NH3-N, being

a tissue toxin, is detoxified to urea by the liver. Urea synthesized in the liver is transported by

the blood, partially reabsorbed in the kidneys, and can be recycled back to the digestive tract

(Stewart and Smith, 2005), contributing to the ruminal NH3-N pool. A large portion of the

dietary proteins and non-protein compounds entering the rumen are degraded by the ruminal

microorganisms to peptides, amino acids, and eventually to NH3-N (Hristov and Jouany,

2005). NH3-N is absorbed into the blood stream through the rumen wall or other sections of

the gastrointestinal tract (Reynolds and Kristensen, 2008) and also contributes to urea

synthesis in the liver.

The mechanisms controlling urea recycling in ruminants are complex and, in spite of

intensive research in the past couple of decades, are not completely understood (Reynolds and

Kristensen, 2008). Recent studies have reemphasized the importance of these mechanisms in

preserving N and potential to provide available N for microbial growth when dietary protein

may be deficient (Lapierre and Lobley, 2001; Reynolds and Kristensen, 2008). The level of

dietary CP is one of the most important factors determining urea recycling rate to the gut and

utilization by the microbes in the rumen (Reynolds and Kristensen, 2008). A series of classic

experiments from the Research Center for Animal Production in Dummerstorf-Rostock have

Ammonia Emission from Animal Manure


demonstrated, for example, that the efficiency of supplemental urea utilization for microbial

protein synthesis in the rumen is sharply decreased (respectively, urinary N losses are

increased) as dietary or plant protein availability increases (Voigt et al., 1984; Piatkowski and

Voigt, 1986). Growing cattle (Wickersham, 2006), or dairy cows (Ruiz et al., 2002) fed lowCP diets have the ability to recycle to the gut virtually all urea synthesized in the liver, with

very little being lost in urine. As pointed out by Reynolds and Kristensen (2008), ruminants

will still excrete N in urine, but it will predominantly be N other than urea. Urea transferred to

the gastrointestinal tract will be utilized for anabolic purposes, i.e. microbial protein

synthesis, at a much greater rate in ruminants fed low-CP diets (Reynolds and Kristensen,

2008). Even at high levels of dietary CP intake, cattle will recycle a significant proportion of

urea to the gut. Gozho et al. (2008) studied urea recycling in dairy cows fed diets with 17 to

17.4% CP and reported urea entry rate to the gastrointestinal tract of around 62%. Urea

utilization for microbial protein synthesis in the rumen, however, was only about 20%.

Current feeding systems for ruminants (NRC, 1996, NRC, 2001) do not account for urea

recycling and likely overestimate the protein (particularly ruminally-degradable protein,

RDP) requirements of the animal. Meta-analysis of a large (1,734 diets) dataset demonstrated

that among several dietary and animal performance variables, dietary CP was the most

important factor determining milk N efficiency in dairy cows (Huhtanen and Hristov, 2009).

Variability in milk yield may explain some of the variability in milk N efficiency when

included in a model with dietary CP, but was insignificant as a stand-alone prediction

variable. Other more recent estimations have shown that increasing dietary CP concentration

with 1%-unit may increase milk protein N yield by approximately 2.8 g d-1, but will result in

35.7 g d-1 dietary N not being utilized for milk protein synthesis (Hristov and Huhtanen,

2008). A major fraction of this unaccounted N will be excreted in urine, which is more

susceptible for leaching and evaporative losses than fecal N (Bussink and Oenema, 1998).

Huhtanen et al. (2008), for example, using a dataset of mainly grass silage-based diets

estimated that 84% of the incremental N intake at constant dry matter (DM) intake is excreted

in urine. Therefore, a better control of ruminal N, particularly NH3-N, metabolism is an

obvious way to achieve an improvement in the efficiency of N utilization by ruminants and to

limit the excretion of nitrogenous compounds resulting in environmental pollution around

animal production areas (Hristov and Jouany, 2005).

Urea is the main constituent of ruminant urine. Bristow et al. (1992), among others,

reported that urea N represented from about 60 to 90% of all urinary N in cattle, with similar

proportions for sheep and goats. Other significant nitrogenous compounds are hippuric acid,

creatinine, and metabolites of purine bases catabolism, such as allantoin, uric acid, xanthine,

and hypoxanthine. Bussink and Oenema (1998) summarized existing literature and gave a

range of urinary urea as proportion of total N of 50 to 90%. In the urine of high-producing

dairy cows, urea represents 60 to 80% or more of the total urinary N (Reynal and Broderick,

2005; Vander Pol et al., 2007) and this proportion gradually increases as dietary CP level and

intake increase (Colmenero and Broderick, 2006). Urea is the main source of NH3 volatilized

from cattle manure (Bussink and Oenema, 1998). These authors indicated that 4 to 41% of the

urinary N may be volatilized, while N volatilization from feces is considerably less, 1 to 13%.

Urea is not volatile, but once mixed with feces; it is rapidly hydrolyzed to NH3-N and carbon

dioxide by the abundant urease activity in fecal matter (Bussink and Oenema, 1998). The

following steps in NH3 volatilization have been identified (Monteny and Erisman, 1998): (1)

urea hydrolysis mediated by urease); (2) dissociation (governed by pH and temperature); and


Pius M. Ndegwa, Alexander N. Hristov and Jactone A. Ogejo

(3) volatilization (which is a function of temperature and air velocity). Miner et al. (2000)

demonstrated the relationship between concentration of NH3-N (as percent of the total

ammoniacal nitrogen, TAN) and pH and temperature. Irrespective of ambient temperature,

less than 1% of TAN is in form of NH3-N at pH 7. As pH increases, however, up to 10, 50,

and 85% of TAN is in NH3-N form at pHs of 8, 9, and 10, respectively; with substantial

differences between 10C and 30C. Air velocity over the liquid manure surface has been

identified as the main factor determining NH3 release rates (Ni, 1999).

A series of experiments were recently conducted to quantify the relative contribution of

urinary vs. fecal N to NH3 volatilization losses from cattle manure (Lee and Hristov, 2010a).

Feces and urine of lactating dairy cows were labeled separately with 15N through labeling of

ruminal microbial protein (see Hristov et al., 2005 for labeling protocol), combined in a 1:1

ratio, and incubated for 10 d in a laboratory scale, closed-chamber system. Ammonia emitted

was captured in an acid trap and analyzed for 15N. As either feces or urine were the only

sources of 15N above background enrichment, the origin of NH3 volatilized during the

incubation could be traced and quantitatively determined. Results from this study are

presented in Figures 6.1 and 6.2. The proportion of NH3 originating from fecal N (Figure 6.1)

was negligible in the first 48 h of the incubation, represented 0.04 ± 0.006 by d 5, and then

gradually increased to 0.11 ± 0.019 of the emitted NH3 by d 10. The proportion curve fitted

well a logistic regression model (adjusted R2 = 0.91; P < 0.001). The proportion of NH3

originating from urinary N represented 0.94 ± 0.027 at 24 h, 0.97 ± 0.002 at 48 h, 0.91 ±

0.004 at 72 h, and gradually decreased to 0.87 ± 0.005 at incubation d 10. The curve fitted

well an exponential decay regression model (adjusted R2 = 0.92; P < 0.001). The average

recovery of NH3 by this approach was 0.95 ± 0.011 for the 10 d of manure incubation. This

study also clearly demonstrated that the main source of NH3 volatilized from cattle manure

during the initial 10 d of storage is urinary N, representing on average 90% of the emitted

NH3. The contribution of fecal N was relatively low, but gradually increased to about 10% by

d 10, as mineralization of fecal N progressed. Using a similar approach, Thomsen (2000)

estimated that urinary N accounted for 79% of the total N losses from sheep manure after 7 d

of composting, but only for 64% at the end of the 86-d storage period. If manure was stored

anaerobically, urinary N accounted for 94% of the total N losses after 28 d and for 68% at 86


Reducing ration CP (Frank et al., 2002) or RDP concentration (Van Duinkerken et al.,

2005) effectively reduces volatile N losses from manure. Metabolizable protein supply has

similar effect. Weiss et al. (2009) reported that increasing dietary metabolizable protein

increased NH3 produced per gram of manure mainly because of increased urinary N excretion

with a significantly smaller contribution of fecal N. Recent studies also show a remarkable

effect of decreasing dietary CP on NH3 emitting potential of dairy manure. In one study, a

replicated Latin square design with 6 ruminally-cannulated cows, 3 diets varying in CP

(HighCP, 15.4; MedCP, 13.4, and LowCP, 12.9% CP, DM basis) and RDP, but having

similar metabolizable protein were fed to lactating dairy cows (Agle et al., 2010). Both

MedCP and LowCP resulted in lower ruminal NH3-N pool size and absolute and relative

excretion of urinary N compared with the HighCP diet. Excretion of fecal N and milk yield

and composition were not affected by diet. As a result of the greater urinary N excretion with

the HighCP diet, cumulative (15 d) NH3 emissions from manure were significantly greater (P

< 0.001) for HighCP compared with MedCP and LowCP (2,278, 1,673, and 1,418 mg,

respectively; Figure 6.3). The rate of NH3 emission was also considerably greater for HighCP


Ammonia Emission from Animal Manure

compared with the low-CP diets (138, 98, and 83 mg NH3 d-1, respectively; P < 0.001). These

observations were confirmed in a follow-up experiment (Lee and Hristov, 2010b) with highproducing dairy cows (average milk yield was 38 kg d-1) fed high (16% CP, HighCP) or lowCP (14% CP, LowCP) diets. Ammonia emission rates were 202 and 132 mg N h-1 for HighCP

and LowCP manure, respectively (P < 0.001). Cumulative NH3 emission was 45% less (P <

0.001; Figure 6.4) for LowCP compared with HighCP manure. Manure produced from these

diets was applied to 61 × 61 × 61 cm lysimeters collected from a Hagerstown silt loam (fine,

mixed, mesic Typic Hapludalf) in order to determine NH3 emissions from soil amended with

manure from low- and high-CP diets. Manure application rate was 9.3 g of N lysimeter-1,

corresponding to a field application rate of 300 kg N ha-1, and was identical for the 2 types of

manure. The HighCP manure had higher N content (4.4 vs. 2.8%, DM basis) and proportion

of NH4+-N and urea-N in total manure N (51.4 vs. 30.5%) than the LowCP manure and as a

result, more LowCP than HighCP manure (2.36 vs. 1.65 kg) was applied to each lysimeter.

After manure application, NH3 emissions were measured using a photoacoustic infrared gas

analyzer at 3, 8, 23, 28, 50 and 100 h. As Figure 6.5 shows, NH3 emission was significantly

greater (P < 0.05) from HighCP- than from LowCP-manure amended soil. The area under the

cumulative (100 h) NH3 emission curve for LowCP was smaller (P < 0.05) than the area for

HighCP manure (56.8 and 114.8 mg NH3 m-2 min-1 × h, respectively). This experiment

clearly demonstrated that manure from dairy cows fed reduced CP diets had decreased NH3

emitting potential and would result in significantly lower NH3 volatilization when applied to

soil, compared to manure from cows fed a high-CP diet.










Predicted (logistic model)









Incubation day

Figure 6.1. Proportion of ammonia originating from fecal N in dairy manure incubated for 10 d in a

closed-chamber system (from Lee and Hristov, 2010a). Symbols are measured (means ± SE) and lines

are predicted values (logistic regression model).


Pius M. Ndegwa, Alexander N. Hristov and Jactone A. Ogejo



Predicted, Exponential decay















Incubation day

Figure 6.2. Proportion of ammonia originating from urinary N in dairy manure incubated for 10 d in a

closed-chamber system (from Lee and Hristov, 2010a). Symbols are measured (means ± SE) and lines

are predicted values (exponential decay regression model).

With high-producing cows lowering dietary CP may, in certain situations, result in

decreased milk yield (Broderick, 2003), which would be unacceptable to most producers and

nutritionists in the field. These performance effects in most cases stem from the complex

interactions of protein with DM and energy intake (Huhtanen and Hristov, 2009). In the study

mentioned above (Lee and Hristov, 2010b), for example, milk yield was significantly reduced

(by 3 kg d-1; P < 0.04) by the low-CP diet. This effect was clearly a result of essential amino

acid deficiency and could be avoided by providing sufficient metabolizable protein, or

supplementing the diet with synthetic, ruminally-protected amino acid limiting milk

production (Broderick et al., 2008).

Cumulative ammonia N loss, mg


HighCP - actual

MedCP - actual

LowCP - actual

HighCP - regression

MedCP - regression

LowCP - regression
















Incubation day

Figure 6.3. Effect of dietary crude protein concentrations on cumulative ammonia losses from dairy

manure (from Agle et al., 2010). Symbols are measured (means ± SE) and lines are predicted values

(linear regression). HighCP, MedCP, and LowCP are diets with crude protein concentration (DM basis)

of 15.4, 13.4, and 12.9%, respectively.


Ammonia Emission from Animal Manure

Cumulative ammonia emission, mg




Lines, P < 0.001




High CP, 16% CP

Low CP, 14% CP










Incubation time, h

Figure 6.4. Effect of dietary crude protein concentrations on cumulative ammonia losses from dairy

manure (from Lee and Hristov, 2010b). HighCP and LowCP are diets with crude protein concentration

(DM basis) of 16 and 14%, respectively.

Ammonia emission rate, mg NH3/m2/min


HighCP, actual

LowCP, actual

HighCP, predicted

LowCP, predicted













Time, h

Figure 6.5. Ammonia volatilization from soil amended with manure from dairy cows fed high(HighCP, 16%), or low-crude protein (LowCP, 14%) diets (from Lee et al., 2010). Symbols are

measured and lines are predicted values (peak, modified Gaussian and exponential decay regression

models, respectively).


Pius M. Ndegwa, Alexander N. Hristov and Jactone A. Ogejo

6.2.2. Ammonia Release Mechanisms

As noted in the previous subsection, livestock manures contain N in both organic and

forms. Excess N in the feed and inefficient utilization of CP or amino acids in diets is the

source of this N in excreted urine and feces. Most of the N (up to 97%) is excreted as urea in

the urine of sheep, cows, and pigs; while the rest is excreted as undigested organic N in the

feces (McCrory and Hobbs, 2001; Varel, 1997; Mobley et al., 1995). Within hours to a few

days, urea is hydrolyzed to NH4+-N in a process catalyzed by the microbial enzyme urease

originating mainly from feces (Beline et al., 1998). In contrast, the microbial breakdown of

organic N in feces into NH4+-N in a process referred to as ammonification or mineralization

requires months or even years to effect. The NH4+-N resulting from either urea hydrolysis or

organic N decomposition, or both, is the one susceptible to volatilization from manure

depending on pH and temperature conditions.

The hydrolysis of urea to NH3-N (or to NH4+-N) in aqueous environments, which is

catalyzed by the enzyme urease (with nickel as the co-factor in the urease active sites) occurs

in two steps (Kaminskaia and Kostic, 1997; Udert et al., 2003; Banini et al., 1999; Todd and

Hausinger, 1989). In the first step (depicted in Equation 6.1), a mole of urea is hydrolyzed

into a mole of NH3-N (or NH4+-N depending on pH conditions) and a mole of the unstable

carbamic acid. The mole of unstable carbamic acid then spontaneously decomposes into

another mole of NH3-N and a mole carbon dioxide in a second step (presented in Equation

6.2). A mole of urea, therefore, produces two moles of NH3-N. There are no documented

cases of uncatalyzed hydrolysis of urea in aqueous solutions (Kaminskais and Kostic, 1997),

which demonstrates the importance of urease in the formation process of NH3-N from urea.

NH2 CO NH2 + H2 O → NH3 + NH2 CO OH


NH2 CO OH → NH3 + CO2


The conversion of organic-N (proteins, amino polysaccharides, and nucleic acids) to

NH3-N or NH4+-N, on the other hand, is mediated by a host of enzymes produced by

heterotrophic microbes (Vavilin et al., 2008; Zhang et al., 2007; Horton et al., 1992). The

process also takes place in two distinct stages. First, extracellular enzymes (e.g. proteases,

peptidases, chitinase, chitobiase, lyzosyme, ribonucleases, deoxyribonucleases, exonucleases,

and endonucleases) break down organic-N polymers into monomers (amino acids, amino

sugars, and nucleic acid). Second, the monomers then pass across the microbial cell

membrane and are further metabolized by intracellular enzymes (e.g. dehydrogenases,

oxidases, and kinases) into NH4+-N (Barracklough, 1997; Barak et al., 1990). Some of the

NH4+-N is incorporated into the microbial biomass in a process referred to as assimilation; the

excess or surplus NH4+-N is released back into the bulk manure. The mineralization of protein

N to NH4+-N, for example, involves: (i) the formation of intermediate amino acid N from

protein N which is catalyzed by proteases and (ii) hydrolysis of this amino acid N to NH4+-N,

which is catalyzed by either amino acid dehydrogenases or amino acid oxidases (Nannipieri

and Eldor, 2009).

Ammonia Emission from Animal Manure




Figure 6.6. Equilibrium between NH4+-N and NH3-N in aqueous solutions as a function of pH and


As mentioned earlier, NH4+-N itself is not volatile but it is amenable to volatilization

once it is dissociates into the NH3-N. In aqueous environments, NH4+-N and NH3-N exist in

equilibrium that is governed by both pH and temperature conditions. At constant temperature,

for example, the pH of manure determines the equilibrium between NH4+-N and NH3-N (see

Equation 6.3). Lowering the pH of manure results in an equilibrium shift that favor the NH 4+N form, which effectively lowers the potential of NH3 volatilization. On the other hand,

raising the pH pushes the equilibrium towards the NH3-N form, which exacerbates NH3

volatilization. In general, NH3 volatilization is directly proportional to the proportion of NH3

in the total ammoniacal nitrogen (TAN = NH4+-N + NH3-N) in the aqueous solutions such as

the manure slurries. The influence of pH and temperature on the dissociation of NH4+-N is

shown in Figure 6.6. The fraction of TAN present as NH3-N increases with increase in the pH

of the manure (Figure 6.6a). For a given pH, on the other hand, the fraction of TAN present as

NH3-N increases with increase in temperature (Figure 6.6b). At pH values lower than 8.3 (at

25C), the proportion of TAN present as NH3-N is less than or equal to 0.5. Decreasing

proportion of NH3-N in liquid manure results in a lower potential of NH3 to volatilize (Figure

6.6a: Sawyer and McCarty, 1978). The greatest change in the proportion of TAN present as

NH3-N occurs between pH 7 and pH 10. Below pH of 7 NH3 volatilization decreases

progressively to about pH 4.5 where there is almost no measurable NH3-N (Ndegwa et al.,

2008; Hartung and Phillips, 1994; Sawyer and McCarty, 1978). The influence of temperature

on ammonium-ammonia equilibrium, on the other hand is shown in Figure 6.6b (Loehr,

1974). Increasing temperature increases dissociation of NH4+-N to NH3-N and thus also

enhances NH3 volatilization.



NH3 + H+


Theoretically, the process of NH3 volatilization involves movement of NH3-N to the

manure surface followed by its subsequent release into the ambient air (Teye and Hautala,

2008; Ni, 1999). A conceptual model of NH3 formation and volatilization is presented in

Figure 6.7. Transfer of NH3-N to the manure surface is achieved through diffusion mass

transfer because of concentration gradient, while the release of NH3 from manure surface to

ambient air is mainly through convective mass transfer (Ni, 1999; Kirk and Nye, 1991; van


Pius M. Ndegwa, Alexander N. Hristov and Jactone A. Ogejo

der Molen et at., 1990; Olesen and Sommer, 1993). The resistance of NH3-N transfer to the

surface is relatively small compared to the resistance of its release into the ambient air. The

latter process is thus more significant to the overall NH3 volatilization process. In general,

NH3 volatilization increases with an increase in: concentration of NH3-N in the manure near

the surface of manure containment, air velocity and turbulence, manure temperature, and

manure pH (Teye and Hautala, 2008; Arogo et al., 1999; Sommer et al., 1991; Olesen and

Sommer, 1993; Vlek and Stumpe, 1978). There are, however, emerging thoughts that the

concentration of NH3-N at the surface of the manure containment where volatilization occurs

may be different from the bulk liquid concentration. De Visscher et al. (2002) reported that,

although suspended solids are involved in the transport of the NH3-N to the liquid interface,

their effect may not be accounted for in the estimation of NH3 volatilization using bulk liquid

concentrations only. Ro et al. (2007) observed that gas bubbles, methane in particular, may

also increase NH3 volatilization via entraining of NH3 from the sludge layer where anaerobic

digestion occurs. The NH3 entrained in the bubbles is also not accounted for in the estimation

of volatilization using bulk liquid concentration.


Techniques for reducing NH3 losses from animal manure may be put into three broad

categories: (i) those that minimize N in the animal manure prior to its excretion, (ii) those that

reduce volatile N species (NH3-N or NH4+-N) in the excreted manure, and (iii) those that

physically contain and treat NH3-N or NH4+-N species after they have already been formed.

The basic principles behind each of these NH3 mitigation strategies are discussed in this

section. An overview of the performance of each technique is presented following the

discussion of the principle or strategy in question.

NH3(g) Release

Free Air Stream

Manure Surface

Convective mass transfer


Diffusion mass transfer

Bulk Manure

[NH4+] + [NH3]

Urea and Organic-N

NH4+-NH3 Equilibrium (pH

and temperature dependent)

Enzymatic and microbial processes

Figure 6.7. A conceptual model of ammonia formation and volatilization from manures.

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