Hatfield J.L., Gandini A. (eds.) Nitrogen in the Environment
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Gaseous Nitrogen Emissions from Livestock Farming Systems 405
sodium (Na), potassium (K), and N in the urine appeared to explain most of the
variation in the volume of urine excreted. Hence, the more salts and N excreted
with urine, the larger the total urine volume excreted. This principle holds for other
animal species as well.
Increasing the volume of urine by increasing the amount of electrolytes in the
diet decreases the concentration of urea in the urine and increases the frequency
of urination and also increases the volume of a single urination. Such multiple and
interactive effects complicate the prediction of the effects of changes in animal ’ s
diet on N excretion and especially on NH
3
volatilization ( Smits et al., 1995 ).
Also the pH of freshly excreted urine mainly depends on the dietary content of
electrolytes ( Figure 5 ), and this principle holds for all animal species. Although the
pH will eventually rise toward alkaline values due to the hydrolysis of urea irrespec-
tive of initial pH, it is the initial pH and the pH buffering capacity of urine which
determines the rate of NH
3
volatilization from urine immediately following urination.
Volatilization of NH
3
from animal manure is closely related to urinary excre-
tion of urea, which is rapidly hydrolyzed by the enzyme urease into NH
4
/NH
3
and
HCO
3
according to:
CO NH H O (NH ) CO NH CO NH NH HCO()
22 2 42 3 4 3
2
34 3
22
↑
0 10203040506070
Calculated urine production (kg/day)
0
10
20
30
40
50
60
70
Observed urine production (kg/day)
Figure 4 . Comparison between measured and calculated urine production by
lactating dairy cows. Calculated values were obtained from regression analy-
sis, according to: urine production (kg/day) 0.134 (0.010) urinary excreted
sodium (g/day) 0.061 (0.005) urinary excreted potassium (g/day) 0.024
(0.007) urinary excreted N (g/day) (R
2
0.898, SE 4.22 kg/day) ( Bannink
et al., 1999 ).
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406 Nitrogen in the Environment
Grazed pastures and the floors in animal houses that are fouled with fecal mat-
ter have a high urease activity and the hydrolysis of urea into NH
3
/NH
4
is gener-
ally completed within a few hours or days. The hydrolysis of urea is associated with
an increase in pH because of the production of (bi)carbonate(s). This combined
with the high pK
a
value (9.15) of the NH
3
/NH
4
equilibrium make that urea pud-
dles have a high potential for NH
3
volatilization.
A low pH of the urine decreases the urease activity and also affects the NH
4
/
NH
3
equilibrium and the NH
3
partial pressure. Ruminants normally consume large
amounts of forage which is relatively rich in cations. As a result, the urine excreted
by ruminants is alkaline ( Figure 5 ) under normal conditions and without salt sup-
plementation. With introduction of acidifying salts in the diet of growing pigs , Cahn
et al. (1998b) established a decrease of the pH of fresh urine by more than two units
and a decrease of the volatilization of NH
3
from slurry by up to 50%. Stimulation
of fermentation in the large intestine also lowers the pH of slurry because it will
contain a higher concentration of VFA as end products of microbial fermentation.
In controlled experiments with growing pigs, Cahn et al. (1998a) obtained a 45%
decrease in NH
3
volatilization from pig slurry when starch-rich tapioca constituting
15% of dietary dry matter was replaced by pressed sugar beet pulp rich in struc-
tural carbohydrates. Although such experimental diets are not representative for the
diets used in practice, it shows that fermentation in the large intestine is primarily
an important determinant for the site of N excretion in pigs and the NH
3
volatiliza-
tion following urination. However, also pH (excretion of VFA) and fecal dry matter
content (volume of feces) may be significantly affected ( Cahn et al., 1998a ;
600 400 200 0 200 400 600
Dietary cation anion difference (meq/kg feed)
5
6
7
8
9
Urine (pH)
Figure 5 . The relationship between dietary cation–anion balance (Na K
–
Cl
–2
S
in meq/kg feed dry matter) and the pH of urine excreted by dairy cows. The follow-
ing relationship was derived by regression of a logistic equation to observed data:
Urine pH 5.72(0.17) 2.57(0.29)/[1exp ( 0.015(0.004) (cation–anion bal-
ance)] R
2
0.692; SE 0.612 ( Bannink and van Vuuren, 1998 ).
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Gaseous Nitrogen Emissions from Livestock Farming Systems 407
Portejoie et al., 2004 ) and an important determinant for NH
3
concentrations in
slurry and consequent NH
3
volatilization rates.
4 . GASEOUS N LOSSES FROM DUNG AND URINE IN PASTURES
Grazed grasslands are aggregations of grazed and ungrazed areas, urine patches,
dung patches, compacted hoof holes, camping areas, and mixtures of these areas. At
the end of the growing season of intensively grazed grassland, there is a mosaic of
fresh and old, and of single, overlapping and possibly mixed urine and dung patches
covering up to 40% of the total area ( Lantinga et al., 1987 ). The spatial distribution
over the field of patches containing fecal and urinary N is known to be very hetero-
geneous ( Afzal and Adams, 1992 ). The urine and dung patches have a much higher
potential for emitting NH
3
, NO, N
2
O, and N
2
into the atmosphere than the urine-
and dung-free areas. As a consequence, the spatial variability of gaseous N losses
in grazed grasslands is extremely high ( Colbourn 1992, 1993 ; Van Cleemput et al.,
1994 ; Velthof et al., 1996a, b ; Oenema et al., 1997 ).
This section reviews literature on gaseous N losses from dung and urine in pas-
tures. It complements a recent review by Bolan et al. (2004). The effects of manure
application on gaseous N emissions from pastures (and arable land) are discussed in
Section 6.
4.1 . Volatilization of NH
3
from Pastures
Most of the published experimental studies on NH
3
volatilization from grazed
pastures date from the 1980s to early 1990s. Later (after 1995) experimental studies
on NH
3
volatilization mainly focus on the effects of manure application techniques
(see Section 6) and on livestock buildings and hard standings and manure storage
facilities (Section 5).
The loss of NH
3
from pastures is related to pasture productivity, grazing intensity,
fertilizer and manure application, and environmental conditions. Total-NH
3
losses
strongly relate to the amount of N in dung and urine deposited by grazing animals,
which ranges from less than 50 kg/ha/year in extensively (nonfertilized) managed pas-
tures to more than 200 kg N/ha/year in intensively managed grasslands ( Whitehead,
2000 ). Ammonia emission factors range from 3–15% of the amount of N excreted
( Ryden et al., 1987a ; Jarvis et al., 1989 ; Bussink 1992, 1994 ), with the median value
close to 7% ( Bouwman et al., 1997 ; Misselbrook et al., 2000 ). Sherlock and Goh
(1984) found a large seasonal variability in emission factors: from 25% in autumn to
12% in winter. The strong seasonal variation has been related to variations in weather
conditions, grazing periods and fertilizer applications ( Denmead et al., 1974 ; Ball and
Ryden, 1984 ; Jarvis and Pain, 1990 ; Bussink, 1992, 1994 ; Petersen et al., 1998a ).
Plantaz (1998) showed that net emission of NH
3
occurred during the grazing season
and a net deposition of NH
3
during the nongrazing winter season. He also showed that
NH
3
emission via the stomata of plants was a way of NH
3
emission from grassland
toward the atmosphere. Similar observations have been made by Harper et al. (1996) .
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408 Nitrogen in the Environment
Enclosure measurements of NH
3
volatilization from single urine patches indi-
cate that NH
3
losses may range from 4% to 52% of the urine N ( Vallis et al., 1982 ;
Sherlock and Goh, 1984 ; Vertregt and Rutgers, 1988 ; Lockyer and Whitehead,
1990 ; Whitehead and Raistrick, 1991, 1993 ; Petersen et al., 1998a ). More than 50%
of the NH
3
volatilization from urine patches occurred during the first few days after
urine deposition. Largest NH
3
losses from urine-treated soil were found at high
urea-N concentrations ( Petersen et al., 1998a, b ), high temperatures ( Lockyer and
Whitehead, 1990 ), relatively dry conditions ( Whitehead and Raistrick, 1991 ), and a
relatively low cation exchange capacity of the soil ( Whitehead and Raistrick, 1993 ).
Whitehead et al. (1989) showed that the composition of urine has a strong influ-
ence on NH
3
volatilization; NH
3
volatilization from the five major components of
urine decreased in the order urea allantoin creatinine creatine hippuric
acid. However, the NH
3
volatilization from a mixture of hippuric acid and urea was
higher than from urea only, probably because hippuric acid increases soil pH. The
NH
3
volatilization from dung pats are (much) smaller than from urine patches and
range from less than 1% ( Ryden et al., 1987a ; Petersen et al., 1998a ) to 13% of the
dung N ( Vertregt and Rutgers, 1988 ).
An integral estimate of the NH
3
loss from grazed pastures, that is, from both the
dung and urine contaminated areas and from the dung-free and urine-free grazed
and ungrazed areas can be obtained from micro-meteorological studies. Table 3
provides an overview of results obtained with micro-meteorological methods. Total
annual NH
3
volatilization from grazed grasslands ranges from 1 to 42 kg N/ha/year
or 3.3% to 14.4% of the total-N excreted. Losses of NH
3
tend to be lower from
sheep-grazed grasslands than from cattle-grazed grasslands, possibly because of
the lower N content in sheep urine and the smaller volume of a single urination
( Bristow et al., 1992 ).
The NH
3
volatilization from grazed grassland is related to N fertilizer input.
Fertilizer N increases herbage production and the N content of the herbage. As
a consequence, more animals can graze the pasture and more N is excreted by
the grazing animals. Further, more N will be excreted via urine when the protein
content of the herbage increases. Misselbrook et al. (2000) derived the following
relationship between NH
3
losses from grazed grasslands and the input of mineral
fertilizer N, using data of measured NH
3
volatilization of complete grazing seasons
in the United Kingdom, The Netherlands, and New Zealand:
NH loss fertilizer N input
3
2 27 0 0683..×
where NH
3
loss is the loss in grams of N per live weight unit per day (1 live
weight unit 500 kg), and fertilizer N input is the fertilizer application rate in kg
N/ha/year. Because the stocking density increases with an increase in fertilizer N
input, the equation indicates that the NH
3
loss increases more than proportionally
with an increase in N input.
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Gaseous Nitrogen Emissions from Livestock Farming Systems 409
4.2 . Volatilization of NO, N
2
O, and N
2
from Pastures
Freshly excreted urine and dung contain energy-rich and chemically reduced car-
bon and nitrogen compounds, which provide substrates for consortia of autotrophic and
heterotrophic bacteria. Upon partial aeration, autotrophic nitrifiers oxidize the
N H
4
mineralized from urine and dung to
NO
2
and subsequently to
NO
3
. The
rate of nitrification depends on the NH
4
and oxygen (O
2
) contents, pH, and tem-
perature. Usually, it takes weeks before all NH
4
in urine patches have been
converted into
NO
2
and
NO
3
. The high osmotic pressure and high NH
3
par-
tial pressure in urine patches may partly inhibit the nitrification process ( Darrah
Table 3 .
Annual emissions of NH
3
from grazed grasslands.
Site
Type of
animal
Fertilizer
input (kg
N/ha/year)
NH
3
-N
emission
(kg ha/year)
NH
3
-N
emission
(% of total
excreted N)
NH
3
-N
emission (%
of urinary N) Reference
UK Beef cattle 210 10 6.7 11.2 Jarvis
et al., 1989
UK Beef cattle 420 25 9.0 12.1 Jarvis
et al., 1989
UK Beef cattle 0 (clover) 7 Jarvis
et al., 1989
UK Sheep 420 9 Jarvis
et al., 1991
UK Sheep 0 4 Jarvis
et al., 1991
UK Sheep 0 (clover) 1 Jarvis
et al., 1991
NL Dairy cows 550 39–42 7.7–8.5 Bussink,
1992
NL Dairy cows 250 4–9 3.3–5.3 Bussink,
1992
NL Dairy cows 400 12–27 6.9–13.9 Bussink,
1994
NL Dairy cows 550 15–33 6.9–14.4 Bussink,
1994
NZ Dairy cows 0 (clover) 15 3.5 Ledgard
et al., 1996
The emitted amount of NH
3
-N is expressed in percent of the total-N or the
urinary-N excreted by the grazing animal. UK, United Kingdom; NL, the
Netherlands; NZ, New Zealand.
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410 Nitrogen in the Environment
et al., 1985 ; Monaghan and Barraclough, 1992 ). Further, nitrite-oxidizing bacteria are
more rapidly inhibited than the NH
3
-oxidizing bacteria. As a consequence, a tempo-
rary accumulation of
NO
2
may develop in urine and dung patches, which promotes
the production and release of NO and N
2
O from the urine and dung ( Monaghan and
Barraclough, 1993 ). Low O
2
concentrations in the soil may also promote the tempo-
rary accumulation of
N
O
2
during nitrification and hence, the release of NO and N
2
O.
Heterotrophic denitrifiers use the
N
O
3
and
NO
2
as electron acceptors; they
chemically reduce these N compounds to NO, N
2
O, and N
2
. The rate of denitrifica-
tion and the release of NO, N
2
O and N
2
into the atmosphere are controlled by (i)
N
O
2
and
NO
3
contents, (ii) availability of organic C as substrate for the het-
erotrophic denitrifiers, (iii) O
2
content, (iv) pH, and (v) temperature ( Tiedje, 1988 ).
Urine and dung are sources of both
N
O
3
and available organic C and, therefore,
the denitrification activity can be very high in urine and dung affected soil ( Ryden,
1986 ). Treading and trampling by grazing animals also contribute to denitrifying
activity because of soil compaction ( Warren et al., 1986 ; Naeth et al., 1990 ). Soil
compaction retards water infiltration rate and O
2
diffusivity in soil, enhancing both
N
2
O productions during nitrification and denitrification activity ( Torbert and Wood,
1992 ; Hansen and Bakken, 1993 ; Van Groenigen et al., 2005a, b ).
Results from experiments dealing with N
2
O emission and denitrification losses
from dung and urine on pastures are shown in Table 4 . The emission of N
2
O from
soils is generally measured using enclosures ( Mosier, 1989 ). Denitrification is gen-
erally determined using the acetylene-inhibition technique with intact soil cores
( Ryden et al., 1987b ), or estimated as the unexplained part of the N balance ( Pakrou
and Dillon, 1995 ; Thompson and Fillery, 1998 ). In some studies
15
N labeling
was used to quantified gaseous N losses (e.g., Monaghan and Barraclough, 1993 ;
Clough et al., 1998 ). The N
2
O emission from dung pats ranges from 0.1% to 0.7%.
However, the amount of mineralizable organic N in dung pats is large, which can
make these pats conducive to N
2
O release via nitrification and denitrification over a
long period ( Yamulki et al., 1998 ). These findings suggest that the data presented so
far underestimate the N
2
O emission from dung, because most measurement periods
have been relatively short ( Table 4 ). Emissions of N
2
O from urine patches range
from 0.0% to 15.5%. This wide range has been attributed to variations in urine
composition, soil type, and environmental conditions, which all can have large
effects ( Sherlock and Goh, 1983 ; Monaghan and Barraclough, 1993 ; Allen et al.,
1996 ; Yamulki et al., 1998 ; Van Groenigen et al., 2005a, b ). Recent results have
shown that the nitrogenous composition of urine, especially hippuric acid content,
may have large effects on N
2
O emissions ( Kool et al., 2006 ). Denitrification losses
from urine affected soils range from less than 1% of the urine N under dry condi-
tions up to 65% of the urine N under moist conditions ( Table 4 ). Monaghan and
Barraclough (1993) showed large losses of N
2
during the first days after urine appli-
cation, which was attributed to increased denitrification as a result of an increased
amount of available carbon in soil, an increased microbial activity and a decreased
oxygen content in the soil.
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Gaseous Nitrogen Emissions from Livestock Farming Systems 411
Table 4 .
Emission of N
2
O from animal dung and urine deposited in grassland; a compilation
of published and unpublished data.
Country
1
Soil Object
Period
(days)
N
2
O
emission
(% of
excreted N)
Denitrifi cation
(% of
excreted N) Reference
UK clay
loam
dung 66–417 0.1–0.7 Yamulki et al.,
1998
G Loess dung 77 0.5 Flessa et al.,
1996
G Sand dung 365 0.4 Poggemann
et al., 1995
NL Sand dung 184 0.7 Velthof, unpub-
lished data
USA clay
loam
urine 300 0.6 Mosier and
Parton, 1985
UK clay
loam
urine 30 1.0–5.0 Monaghan and
Barraclough,
1993
UK clay
loam
urine 60–417 0.1–1.4 Yamulki et al.,
1998
NZ clay
loam
urine 12 4.3 Williams
et al., 1998
UK Coarse
silt
urine 42 7.0 Williams
et al., 1999
B sandy
loam
urine 19–35 0.1–2.4 Vermoesen
et al., 1997
USA sandy
loam
urine 530 5.0 Mosier et al.,
1998b
NZ sandy
loam
urine 112 1.0 Clough et al.,
1998
NZ Peat urine 100 1.0 Clough et al.,
1996
NZ Peat urine 112 1.9 Clough et al.,
1998
NZ Peat urine 80 0.3–0.6 De Klein et al.,
2003
UK Peat urine 16 0.4–0.5 Colbourn, 1992
NZ silt
loam
urine 100 1.5–3.0 Clough et al.,
1996
(Continued)
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412 Nitrogen in the Environment
Table 4. (Continued)
Country
1
Soil Object
Period
(days)
N
2
O
emission
(% of
excreted N)
Denitrifi cation
(% of
excreted N) Reference
NZ silt
loam
urine 112 0.8 Clough et al.,
1998
NZ silt
loam
urine 21 0.5–6.4 Clough et al.,
2003b
NZ silt
loam
urine 60 0.8 Clough et al.,
2003a
NZ silt
loam
urine 150–351 0.4–3.7 De Klein et al.,
2003
UK silt
loam
urine 16 0.7–2.0 Colbourn,
1992
G sandy silt
loam
urine 357 0.3–4.2 Anger et al.,
2003
NZ stony silt
loam
urine 231 0.5 De Klein et al.,
2003
NZ – urine 42 0.5 Sherlock and
Goh, 1983
G Loess urine 77 3.8 Flessa et al.,
1996
NL Clay urine 28 0.5 Velthof and
Oenema, 1994
UK Clay urine 38 12.7–15.5
3
Lovell and
Jarvis 1996
NZ Clay urine 112 1.9 Clough et al.,
1998
G Sand urine 365 0.4–1.3 Poggemann
et al., 1995
AUS Sand urine 28 0.0 Bronson et al.,
1999
NL Sand urine 103 0.2–2.3 Van Groenigen
et al., 2005a
NL Sand urine 35 0.0–10.8 Van Groenigen
et al., 2005b
UK clay loam grazing 7 8.0 Velthof et al.,
1996a
NL Sand grazing 730 1.5 Velthof et al.,
1996c
(Continued)
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Gaseous Nitrogen Emissions from Livestock Farming Systems 413
Table 4. (Continued)
Country
1
Soil Object
Period
(days)
N
2
O
emission
(% of
excreted N)
Denitrifi cation
(% of
excreted N) Reference
NL Clay grazing 730 3.3 Velthof et al.,
1996c
NL peat I grazing 730 2.3–9.8 Velthof et al.,
1996c
NZ silt
loam
grazing 730 0.2 Carran et al.,
1995
NZ silt
loam
grazing 730 1.0 Carran et al.,
1995
NL Sand urine 14 18.0
3
De Klein and
Logtestijn, ‘ 94
NL Peat urine 31 2.2 Koops et al.,
1997
UK clay
loam
urine 30 30–65 Monaghan and
Barraclough,
1993
UK silt
loam
urine 40 0.6–3.2 Colbourn,
1992
NZ Clay urine 31 0.33
2
Clough and
Ledgard, 1997
NZ Peat urine 31 2.35
2
Clough and
Ledgard, 1997
NZ sandy
loam
urine 31 1.35
2
Clough and
Ledgard, 1997
NZ silt
loam
urine 31 0.65 Clough and
Ledgard, 1997
The emitted amount of N
2
O-N is expressed in percent of the amount of N excreted
by the grazing animal.
1
B, Belgium; G, Germany; NL, the Netherlands; NZ, New Zealand; UK, United
Kingdom; AUS, Australia.
2
N
2
loss N
2
O loss.
3
Emissions may have increased due to horizontal diffusion out of soil cores.
Soil compaction in the field increased urine-derived N
2
O emissions from 1.3% to
2.92% ( Van Groenigen et al., 2005b ) and from 0.9% to 4.9% in a laboratory study
( Van Groenigen et al., 2005a ). Combined application with dung increased urine-
derived N
2
O emissions from 0.9% to 7.9% in the laboratory and from 1.6% to 2.82%
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414 Nitrogen in the Environment
in the field. Avoiding combinations of urine patches, dung patches, and compaction
therefore seems to be important for decreasing N
2
O emissions. Grazing-derived N
2
O
emissions range from 0.2% of total excreted N for a silty loam in New Zealand to 9.8%
of the total excreted N for intensively managed grassland on peat soil in the Netherlands
( Table 4 ). Measured denitrification losses from grazed grasslands range from 3 to 19 k g
N/ha/year in New Zealand ( Ruz-Jerez et al., 1994 ; Ledgard et al., 1996 ), 3 to 108 k g N /
ha/year in the Netherlands ( De Klein and Logtestijn, 1994 ; Koops et al., 1996, 1997 ),
and 8 to more than 100 kg N/ha/year in the United Kingdom ( Garret et al., 1992 ).
However, it is not always clear which fraction of the denitrification losses comes from
fertilizer N, biological N fixation, urine N, and dung N. Ryden (1986) showed that
denitrification losses were 1.7 to 6 times higher from grazed grasslands than from cut
grassland with a similar N input. High denitrification rates from grazed grasslands are
found at the end of the grazing season, when N uptake by the grass ceases and soil
water and mineral N contents and soil temperature are still at an elevated level.
Only few studies have quantified the NO emission from urine and dung
patches. Studies of Williams et al. (1998) and Lovell and Jarvis (1996) indicate that
the emission of N
2
O-N is higher than that of NO-N from urine treated soil; the ratio
between NO-N and N
2
O-N soil ranged from 0.001 to 0.011 in laboratory studies
and from 0.005 to 0.263 in a field study. Also Colbourn et al. (1987) showed that
NO-N losses from urine patches are small. Bronson et al. (1999) reported a ratio
of 10, but this was largely related to very low N
2
O emissions rather than high NO
emissions. The NO in urine patches is suggested to be mainly produced by nitrifiers
( Colbourn et al., 1987 ; Williams et al., 1998 ; Bronson et al., 1999 ).
5 . GASEOUS N LOSSES FROM STABLES AND MANURE STORAGE
SYSTEMS
Housed animals deposit feces and urine in the housing system on soil, litter,
concrete floors, or slatted floors. There is a wide variety of animal housing systems,
ranging from simple shelters where animals find protection against sun, rain, cold,
and or wind to climate controlled and mechanically vented housing systems, for
example, poultry in battery cages and finishing hogs.
Animal manure collected in housing systems has to be stored for some time
inside or outside the housing system until timely spreading of the manure on the field,
that is, during the growing season when the crop will be able to utilize the plant nutri-
ents. In animal housing systems with (partially) slatted floors, the urine and feces are
mixed and stored as slurry in pits and channels underneath the slats. When the stor-
age capacity inside the housing system is small, slurry will be stored outside in silos,
tanks, and lagoons. In tie stalls with litter, feces mixed with litter and liquid manure
are collected daily and stored outside in separate storage systems. In deep litter sys-
tems in organic farming, feces and urine are absorbed by straw and compacted by
the loose housed animals. The stacked manure will be removed from the deep litter
system only two to three times a year and transported to a manure heap outside for
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