Aust. J. Soil Res., 1996, 34, 789-802
Ammonia volatilisation from soil irrigated with urban sewage
effluent
C. J. SmithA, J. R. ~ r e n e and
~ ~ W.
, J. BondA
A
Division of Soils, CSIRO, P O Box 639, Canberra, ACT 2601.
Division of Plant Industry, CSIRO, P O Box 1600, Canberra, ACT 2601.
Abstract
Losses of ammonia (NH3) following sewage effluent irrigation of pasture were measured under
different climatic conditions a t Wagga Wagga, New South Wales. Ammonia volatilisation was
measured by the micrometeorological mass balance technique using 2 different passive samplers,
and by an indirect technique based on the measurements of ammoniacal-N ( N H ~ S N H ~ )
concentration, pH, and temperature of the soil solution in the 0-3 mm soil layer, and wind
speed a t 1a2 m above the soil surface. Maximal NH3 emission rates were measured directly
following the effluent-irrigation. There was reasonable agreement between the 2 different
passive gas samplers used to measure NH3 volatilisation. The NH3 volatilised was well related
to the product of wind speed and the equilibrium ammonia concentration (calculated from the
soil solution measurements) as was found in other studies. In addition, NH3 flux density was
strongly related to evaporation; that is, when the water (effluent) evaporated NH3 was lost t o
the atmosphere. Under high evaporative conditions, a maximum of 24% of the ammoniacal-N
in the effluent was lost by volatilisation within 2 days of application.
Additional keywords: nitrogen loss, micrometeorology, gas exchange.
Introduction
Ammonia (NH3) gas is emitted into the atmosphere following the application
of alkaline-producing nitrogenous fertilisers (e.g. urea) or from surface-applied
animal wastes (Hoff et al. 1981; Smith et al. 1988; Schilke-Gartley and Sims 1993;
Gordon and Schuepp 1994; Sherlock et al. 1995). Volatilisation rates are affected
, the
by ( 2 ) the hydrolysis of the organic N compounds to ammonium ( N H ~ )(ii)
exchange between N H in
~ solution and cation exchange sites in the soil, and
~ NH3 in the soil solution, and NH3 in the
(iii) the equilibrium between N H and
gaseous phase (Leuning et al. 1984; Freney et al. 1985; Sherlock and Goh 1985;
Rachpal-Singh and Nye 1986; De Datta et al. 1989; Clay et al. 1990). Both the
hydrolysis and equilibrium reactions are temperature-dependent. Models of NH3
volatilisation (Sherlock and Goh 1985; Rachpal-Singh and Nye 1986; Jayaweera
and Mikkelsen 1990a, 1990b; Jayaweera et al. 1990) predict a diurnal cycle of
volatilisation resulting from the diurnal changes in soil temperature. The highest
rates predicted correspond to maximum soil temperatures. Similarly, wind speed
has a major impact on NH3 volatilisation, with maximum emission corresponding
to the period with the highest wind speed.
Ammonia volatilisation from the surface application of wastes has resulted in
losses ranging from 10 to 99% of the applied N, depending on the edaphic and
environmental factors experienced following application (Beauchamp et al. 1982;
Schilke-Gartley and Sims 1993). Volatilisation from urine or animal dung patches
C. J. Smith et al.
of grazing animals has long been recognised as an important pathway for N loss
from the soil-plant ecosystem (Denmead et al. 1974; Ball et al. 1979; Sherlock
and Goh 1985; Sherlock et al. 1989). Because of the alkaline characteristics of
sewage effluent (Feigin e t al. 1991), NH3 volatilisation may be an important
pathway for N loss when the material is applied to land surfaces. Volatilisation
models (Sherlock and Goh 1985) predict considerable NH3 volatilisation following
irrigation with sewage effluent, but few data are available to validate these
predictions.
This paper reports on the measurement of NH3 volatilisation following the
irrigation of soil with urban sewage effluent. Ammonia volatilisation was measured
with 2 different passive samplers, and an indirect method in which the flux was
calculated from measurements of ammoniacal-N in the 0-3 mm soil layer, surface
pH, temperature of the soil surface, and wind speed measured at 1 . 2 m above
the soil surface. The results from the indirect method were compared with the
micrometeorological method for measuring NH3 loss (Leuning e t al. 1985).
Material and methods
Field site and efluent chemistry
The field experiment was conducted adjacent to the sewage treatment works for the Royal
Australian Air Force a t Forest Hill, approximately 15 km east of Wagga Wagga, New South
Wales. The soils a t the site are classified as a Red Chromosol (Isbell 1993), previously known
as a red podzolic (Stace et al. 1968) or Rhodoxeralf (Soil Survey Staff 1992), a Red Kandosol
(red earth or Ustorthent), and an intergrade between the 2 soils. Some properties of the
surface soil (0-0.05 m) are pH(1: 5 soil :water) 7.3; pH(1:5 soil :0 . 1 M BaCl2) 5.3; cation
exchange capacity (CEC) 10 cmol,/kg; total N 1.49 g/kg; and total C 22.8 g/kg. The area
had been cultivated and sown to wheat (Triticum aestivum) from 1979 to 1981, followed by an
improved pasture phase. The pasture was top-dressed with single superphosphate (150 kg/ha)
in 1988 and with Calphos (100 kg/ha) in 1990.
Secondary treated effluent was obtained from the sewage treatment works. Treatment
included initial filtering to remove grit and processing in an oxidation pond in which oxygen,
produced by surface aeration and algae through photosynthesis, is used by bacteria to decompose
organic matter. Chemical properties of the effluent were pH 8.3; EC 1.26&0.04dS/m;
ammoniacal N 1 1 . 8 3 f 0 . 3 mg/L; NO:-N, 1.24~0.07mg/L; Kjeldahl N, 19 mg/L.
Ammonia volatilisation
Ammonia volatilisation from a 25-m-radius circular area was measured in 2 experiments
following the application of sewage effluent by sprinkler irrigation in December 1994 and
January 1995. A total of 5 l f 7 mm of effluent was applied on 6 December 1994 and 41k5 mm
was applied on 17 January 1995 between 0100 and 0700 hours. Measurements of NH3
volatilisation commenced immediately after irrigation and continued for 4 days. In January,
the NH3 fluxes were low initially and became negligible following 5 mm of rainfall between
1800 and 0800 hours (overnight) on 18 January. On 18 January between 0900 and 1000 hours,
urea was broadcast evenly by hand onto the circular area a t a rate equivalent to 80 kg N/ha
to provide conditions likely to yield high NH3 fluxes.
The vertical flux density of NH3 was determined by the procedure described in Smith et
al. (1988) and Sherlock et al. (1995), using the ammonia sampler developed by Leuning et al.
(1985). Samplers were placed on a mast at the centre of the circular area, 0.2, 0.4, 0.8, 1.2,
2.4, and 3.6 m above the soil surface, and changed 3 times daily. Emission was measured for
the periods 0800-1300 hours, 1300-1800 hours, and 1800-0800 hours (overnight). Background
measurements were made with samplers placed on a mast located a t the upwind side of the
treated area.
In addition, NH3 flux was determined using the passive samplers (Ferm tubes) described
by Schjoerring et al. (1992) and manufactured by Mikrolab Aarhus A/S, Axel Kiers Vej 34,
Ammonia volatilisation from effluent-irrigated soil
DK-8270 Hoejdjerg, Denmark. The passive samplers were coated with oxalic acid (Schjoerring
et al. 1992) and mounted at 0.2, 0.4, 0.8, 1.2, and 2 . 4 m above the soil surface on 4 masts
placed at the corners of the largest square area that could be accommodated within the
circular area. At each height, 2 passive samplers were mounted, one with the stainless steel
disc facing towards the treated area, and the other with the disc facing away from the treated
area (Schjoerring et al. 1992). Ammonia emitted from the circular area was collected through
the open end of one sampler and through the stainless steel disc of the other sampler. The
samplers were changed at 24-h intervals and returned to a mobile laboratory. High purity
deionised water (3 mL) was added to each tube in order to dissolve the ammonium oxalate on
the internal surface. The ammonium concentration of the resulting solution was determined
by measuring the emerald-green colour (660 nm) formed by reaction of ammonia with sodium
salicylate, sodium nitroprusside, and sodium hypochlorite at pH 12.8-13.0 (Technicon Traacs
800, Method No. 825-87T). The Technicon procedure was followed except the buffer was
made by dissolving 58 g NaOH, 20 g ethylenediaminetetraacetic acid (disodium salt), 50 g
sodium potassium tartrate, and 2 mL acetone in water. The solution was cooled to room
temperature and diluted to 1 L. The flux density of NH3 was calculated using the equations
given by Schjoerring et al. (1992).
Indirect procedure for calculating NHaemission
The driving forces for the volatilisation of NH3 are the difference between the NH3
concentration in equilibrium with the soil solution and that in the atmosphere, and wind
speed (Freney et al. 1985; Sherlock et al. 1995). Therefore, the calculated vertical NH3 flux
(F) is given by
where k is an exchange coefficient, u, is the mean wind speed at the reference height (z),
p, is the NH3 concentration a t height (z), and p, is the equilibrium NH3 concentration a t
the soil surface. Denmead (1983) and Freney et al. (1985) have shown that, for an actively
volatilising surface, p, is very much less than p,, thus p. can be set to zero in equation 1.
The calculation of p, requires measurements of the solution ammoniacal-N concentration, pH,
and temperature T ( O K ) and is obtained from the following equations (Leuning et al. 1984;
Freney et al. 1985):
Because of the difficulty of extracting the soil solution from the surface layer, we determined
the surface soil pH with a flat surface electrode, surface soil temperature with thermocouples,
and soil moisture and extractable NH;-N in the 0-3 mm layer by standard methods (Smith
et al. 1988).
Soil temperature was measured a t the surface and at 2.5, 5.0, 7.5, and 10 cm depths
within the soil profile a t 2 sites within the treated area. The surface soil temperature sensors
were installed by pushing a steel needle upwards into the vertical soil face of the small pit
used for the installation of the thermocouples, at an angle of about 30' to the horizontal.
5
The probes were pushed into the preformed holes so that the tip of the probe was ~ 0 . mm
below the soil surface. The other probes were installed horizontally into the vertical soil face
of the installation pit.
At the beginning and end of each NH3 sampling period, a flat-surface electrode and
portable pH meter were used to measure the surface soil pH a t 8 locations upwind of the
central NH3 sampling mast. These values were converted to [H+] concentrations to enable
the calculation of the average pH for a particular sampling period.
Two small soil cores (22 mm i.d.) were taken a t the 8 locations, and immediately separated
into 0-3, 3-7, and 10-20 mm depth sections. In December, only the 0-3 mm depth section
was taken. The samples were returned immediately to a field laboratory a t the site and
C. J. Smith et al.
extracted with 20 mL of 2 M KCl. The suspensions were filtered and the ammonium and
nitrate concentrations of the filtrate were determined using a segmented flow analyser (Alpkem
1992). Soil moisture was determined on a second core taken a t each location by drying the
samples a t 105OC. Mean wind speeds were measured in duplicate with cup anemometers
positioned 1.2 m above the soil surface.
-
3-
2z : n
m
-0
2-
-
-
I"
-
z
-:
-
Fig. 1. Cumulative NH3 loss measured with
the samplers developed by Leuning et al.
(1985) (e),and the samplers and procedures
described by Schjoerring et al. (1992) (a).
(a) Effluent applied in December, (b) effluent
applied in January, and (c) urea applied in
January.
-
0
C
-
1-
m
3
5
1
I
O O
17Jan95
18Jan95
I
I
19Jan95
I
I
20Jan95
Batch isotherms
Batch exchange isotherms for N H were
~ determined in solutions having a constant positive
charge concentration of 30 mmol,/L. Calcium chloride was used in the solution t o maintain
the charge because Ca was the dominant exchangeable cation. A total of 24 different solutions
with NH; concentrations ranging from 0 to 30 mmol,/L were used and measurements for each
N H t concentration were carried out on duplicate soil samples. Each solution (20 mL) was
equilibrated with 2 g of soil by shaking for 1 h on an end-over-end shaker. After equilibration,
the suspension was centrifuged and the supernatant analysed for NH:.
Soluble N H was
~
removed from the soil ~emainingin the centrifuge tube by shaking for 30 min with 20 mL of
70% ethanol (w/w), centrifuging at 2000 G for 10 min, and discarding the supernatant. The
soil was then extracted twice by shaking for 1 h with 20 mL of 2 M KCl. Exchangeable N H ~
was determined on all the 2 M KC1 extracts and summed. The adequacy of the 1 h equilibration
time was confirmed by shaking a series of tubes using the above procedure for 1, 2, and 3 h.
Ammonia volatilisation from effluent-irrigated soil
Evaporation
Evaporation of water from the circular area was measured using PVC cylinders, 0.154 m
i.d. by 0.285 m long, carefully pushed into the soil so that the tops of the cylinders were
level with the surface (Leuning et al. 1994). The cylinders containing the soil (lysimeters)
were removed and a plastic base was attached to the bottom. Each lysimeter was placed in
a plastic bag that was sealed to the outside of the lysimeter with tape. An oversize sheet
metal cylinder (NO. 17 m i.d.; 1 mm wall thickness) was inserted into each cavity, the soil was
excavated from within the cylinder, and a plastic sheet was placed on the floor of the cylinder
to prevent soil adhering to the lysimeter when it was returned to the field. Plants within the
cylinders were sprayed with glyphosate a t the beginning of the measurement period in order
t o minimise water loss by Paterson's curse (E. lycopsis L.). The lysimeters were weighed, t o
a precision of 0 . 1 g, a t 0800, 1300, and 1800 hours each day. A new set of 2 lysimeters was
extracted each day, or after rainfall, to minimise the effect of extraction of water by the plant
roots in the undisturbed soil.
Fig. 2. Ammonia flux density (histograms)
and cumulative ammonia loss ( 0 )following
the application of effluent and urea t o
pasture in December and January. The
ammonia flux density was measured using
the sampler developed by Leuning et al.
(1985). (a) Effluent applied in December,
(b) effluent applied in January, and (c)
urea applied in January.
Results and discussion
The exposure period chosen for the study with the Ferm tubes was 24 h because
of the low flux density of NH3 and the small amount of NH3 collected by these
C. J. Smith et al.
tubes per hour (Schjoerring 1995). Although we were unable to measure the NH3
flux with the Ferm tubes at periods <24 h, there was good agreement between the
rates of NH3 emission measured by the 2 micrometeorological methods (Fig. 1).
Lower flux densities can be measured with the ammonia sampler developed by
Leuning e t al. (1985), and thus, shorter exposure periods can be used. The longer
the period of exposure, the greater the risk of wetting of the internal surface of
the Ferm tubes during periods of heavy rainfall in combination with high winds
(Schjoerring 1995). Furthermore, the need for the 24-h exposure period under
conditions of low NH3 volatilisation precludes any study of the diurnal cycle.
10
6 Dec 94
10
1
7 Dec 94
8 Dec 94
9 Dec 94
I
I
I
I
17 Jan 95
18Jan95
19Jan 95
20Jan95
Fig. 3. Surface soil
temperatures a t the site
during the experimental
periods.
In both experiments, NH3 was emitted at the maximum rate directly following
irrigation. Rain (11 mm) fell between 1800 hours on 7 December and 1300
hours on 8 December and this was sufficient to reduce the NH3 flux to <0.2 pg
N/m2.s (Fig. 2). The NH3 flux remained at this low level for the rest of the
experiment. In both experiments, the NH3 fluxes showed the typical pattern of
emission observed following application of urea or urine (Denmead et al. 1974;
Hoff et al. 1981; Freney e t al. 1983; Ryden and McNeill 1984). The rate of
NH3 emission was always greater during the day than at night. Ammonia fluxes
were higher during the December experiment than during January (Fig. 2) and
this is most likely due to the higher soil temperatures in December (Fig. 3). In
December the soil temperature reached a maximum of about 38°C during the
day, whereas in January the maximum temperature measured on the first few
days after effluent irrigation was ~ 3 0 ° C(Fig. 3). Temperature directly affects
the equilibrium between NH: and NH3 in the soil solution, and thus has a major
influence on NH3 emission.
Ammonia volatilisation from effluent-irrigated soil
Following similar applications of NH:-N
in December and January (6 and
4 . 8 kg N/ha), the loss of N through volatilisation was about 3 times higher in
the December experiments (24%) than in January (8.5%).
In the January effluent experiment, the NH3 fluxes were negligible after 5 mm
of rain fell between 1800 and 0800 hours on 18 January. Urea was applied 1 day
after effluent irrigation. The soil was considered to be sufficiently wet t o initiate
the dissolution of the urea granules and begin hydrolysis of urea. Ammonia
fluxes were larger following the application of urea than those measured after
effluent irrigation (Fig. 2). Emissions increased with time to a maximum value
of approximately 20 mg N/m2 . s measured 24 h after the urea application. The
cumulative NH3 loss over a 48-h period after application amounted to 15 kg N/ha
and corresponded to ~ 1 9 %
of the urea-N applied. Rain started at about 0500
hours on 20 January with a total of 48 mm being measured at the experimental
site. This amount of rainfall would be more than adequate to leach any undissolved
urea and soluble ammoniacal-N from the surface soil (Khanna 1981).
14
-
12
-
10
-
Dec94
%
--5
16
14
+
12
-
10
-
'
94
I
2
-
-
Dec94
94
I
I
:.I:j:::::
4
2
0
-
Fig. 4. Evaporation (0, B) and rainfall
(arrows) a t the experimental site after the
application of sewage effluent.
r.5 -
-
-
l
l
l
I
I
l
_
17Jan9518Jan95 19Jan9520Jan95
Evaporation of water from the soil was significantly greater during December
than in January (Fig. 4). Approximately 5.5 mm of water evaporated from the
soil during the first 24 h in December, whereas only 2.2 mm was lost during the
equivalent period in January. The greater evaporation measured in December than
in January was associated with a larger amount of NH3 being lost in December.
These results suggest that the rate of NH3 volatilisation is strongly influenced
by the rate of water loss from the soil surface following irrigation. Previously,
Denmead et al. (1974) and Freney et al. (1992) showed strong relationships
between the amount of NH3 loss and the amount of evaporation. Freney et al.
(1992) showed that dew which had accumulated on the sugar cane trash during
C. J. Smith et al.
the night controlled the hydrolysis of broadcast urea and thus the amount of
NH3 present for the volatilisation process and that no ammonia was emitted
until evaporation commenced.
100
10
7
1
6 Dec 94
7 Dec 04
8 Dec 94
9 Dec 94
Fig. 5. The pH (+) in the 0-3 mm soil
layer, and extractable ammonium (0)
and nitrate (0)in the 0-3, 3-10, and
10-20 mm layers after application of
effluent and urea (80 kg N/ha).
Since both the rates of evaporation of water from the soil surface and NH3
volatilisation are a function of the energy balance at the soil surface, there is a
strong relationship between the vertical NH3 flux and the evaporation of water.
Under effluent irrigation, the potential amount of NH3 present for volatilisation
will be controlled by the ammoniacal-N concentration in the effluent and the
surface soil pH, and temperature (equations 2 and 3). The pH of the soil surface
was at its maximum value directly after irrigation and was similar to that of the
effluent (Fig. 5). In the days following effluent irrigation, the pH declined to a
value of 6 . 5 . A similar trend in pH was observed in the January experiment.
Immediately after the addition of urea, the pH declined during the day, but
increased during the night. The decline in pH may be attributed to the loss of
NH3 from the soil solution and nitrification, whereas the increase in pH may be
due to hydrolysis of the urea.
Ammonia volatilisation from effluent-irrigated soil
Maximal KC1-extractable ammoniacal-N concentrations in the surface 0-3 mm
were measured directly after irrigation and decreased with time (Fig. 5). There
was a corresponding increase in the KC1-extractable NO; concentrations in all
soil layers for the January 1995 sampling, which showed that the ammoniacal-N
applied in the effluent was rapidly oxidised. The rapid decrease in the NO;
concentrations between 1800 and 1300 hours on December 8, and between 1800
and 0900 hours on January 20, was mainly due to NO; leaching following rainfalls
of 10 and 15 mm, respectively.
Fig. 6. Exchange isotherm for ammonium in
the 0-50 mm layer of soil. The solid line is
the best-fit line defined by equation 4. Data
shown in ( b ) are for the low-concentration
region of (a).
NH,' concentration (rnmolc IL)
The ammoniacal-N concentration in the soil solution was estimated from the
values, using the N H ~ - Nabsorption isotherm shown
KC1-extractable
in Fig. 6. The conventional single-parameter isotherms did not fit the data,
but a good fit to the data was obtained using the 2-parameter relationship of
Rothmund-Kormfeld (Bond and Phillips 1990). The isotherm (solid line) is given by
NHZ-N
C. J. Smith et al.
where S1 is N H sorbed
~
(mmol,/kg of soil)/CEC, and C' is ammoniacal-N in
solution (C, mmol,/L)/total positive charge in solution (TC, 30 mmol,/L). The
cation exchange capacity (CEC) of the 0-10 cm soil layer was 100 mmol,/kg.
The fitted values for kl and n were 0.0790 (f0.0021) and 0.4786 (f0.006),
respectively.
Ammoniacal-N concentration in the soil solution was calculated from the 2 M
KC1-extractable NH4+-N ([NH~$-N]~,~,~)
using the equation
where M,/M, is moisture content of the soil extracted. The above equation
partitions the ammoniacal-N between the sorbed and solution phases when the
2 M KC1-extractable
was less than the CEC of the soil. In this study, we
restricted the use of the equation to NHZ-N values <30 mmol,/kg (equivalent to
420 mg N/kg). When the total NHZ-N values exceeded this value, we assumed
that the amount sorbed was insignificant compared with that in solution, which
is the case following the addition of urea-N.
NHZ-N
Fig. 7. Mean wind speeds a t the site during
the experimental periods.
The factors pH, temperature, and ammoniacal-N concentration in the soil
solution were used to calculate the equilibrium NH3 concentration (po) in solution
at the soil surface, using equations 2 and 3, which was multiplied by the
mean windspeed for the exposure period (Fig. 7) to calculate the ammonia flux
(equation 1). The relationship between these calculated flux values and the
measured NH3 flux values is shown in Fig. 8, along with the regression lines for
the relationships between ammonia flux density and u1.2 PO reported by Sherlock
et al. (1995). In contrast to the results reported by Sherlock et al. (1995), poor
linear relationships were found for our data sets. The high value for ul.apo,
Ammonia volatilisation from effluent-irrigated soil
marked 'A', associated with a low flux is most likely due to the inclusion of
undissolved urea granules in the 0-3 mm soil sample. In this study, the urea
granules were small (<1 mm diameter), and it was impossible to identify them
after application t o the soil surface. Sherlock et al. (1995) reported a value for
the exchange coefficient, k, of 6.9e - 05 following the application of urea to the
soil surface, and 9.0e -05 following the application of synthetic urine, by using
a linear regression forced through the origin. Regression analyses of our sewage
effluent and urea data (excluding point A), with the intercept forced through
zero, gave a k value of 2.74e - 05 (*4.4e - 06; P < 0.05; r2 = 0.65). When
the linear regressions were fitted individually to the effluent data sets, the r2
values for the relationships were 0.54 for the December experiment, and 0.95 for
January; the k values were 7. l l e - 04 (Jzl.77e - 04) and 8.03e - 05 (Jz7.le - 06)
for December and January, respectively. These results confirm the findings of
De Datta et al. (1989), that the exchange coefficient, k, is not constant for all
times. It would therefore be necessary to determine k for each site and time, as
was recommended for urea-fertilised upland crops, urine-affected pastures, and
flooded soils (Sherlock et al. 1995; Freney et al. 1985; De Datta et al. 1989).
Fig. 8. Relationship between measured NH3
flux, and the product of average wind speed a t
1 . 2 m above the soil surface and the calculated
equilibrium NH3 gas concentration a t the
soil surface (p,). Symbols represent sewage
effluent, December experiment (a);sewage
effluent, January experiment (m); and urea,
January (A).
C. J. Smith et al.
Fig. 9. Relationship between measured NH3
flux density, and the water flux density
following the application of effluent.
Water flux density (m31m2.sec)
A strong linear relationship between NH3 flux density and evaporation was
observed for the effluent data sets (Fig. 9). The intercept of the linear regression
was not significantly different from zero, and therefore a zero intercept model
was fitted t o the data. The relationship obtained for the combined sets was
where W is the evaporation of water (m3/m2. s). These data suggest that NH3
volatilisation from effluent-irrigated soil is controlled by the evaporation of 'free'
effluent water at the soil surface. Ammoniacal-N in this fraction would not be
accurately measured by the soil sampling techniques used, which may explain
the poor relationship between NH3 flux and ul . 2 p o The slope of the regression
(equation 6) represents the effective concentration of the ammoniacal-N in the
effluent (pg N/m3 of solution) and is calculated to be 10.2 (f1.26) mg NIL.
This value is consistent with the ammoniacal-N concentration ( ~ lmg
l N/L) that
was measured in the effluent. Thus, for the circumstances of our experiments,
NH3 volatilisation during the first 24 h was controlled by the rate of evaporation
of water and the ammoniacal-N concentration in the effluent.
Conclusions
The results show that there was reasonable agreement between 2 different
passive gas samplers used t o measure NH3 volatilisation. However, we were
unable to measure the NH3 flux with the Ferm tubes at periods of <24 h. In
contrast, lower NH3 flux densities can be measured with the passive gas sampler
developed by Leuning et al. (1985).
Ammonia loss from sewage effluent could be calculated from the exchange
coefficient, k, and measurements of surface soil ammoniacal-N, pH, and temperature,
concurrently with wind speed at the reference height of 1a2 m. It is not possible
to conclude that the magnitude of k would be constant for all sites and times, and
as recommended by others (Freney et al. 1985; De Datta et al. 1989; Sherlock et
al. 1995), the value needs to be determined for each site. A better relationship
was observed between NH3 flux density and evaporation. This may be because
Ammonia volatilisation from effluent-irrigated soil
NH3 volatilisation predominantly occurs in the 24 h following effluent-irrigation
when 'free' effluent water would be evaporating from soil and plant surfaces. In
the absence of NH3 flux measurements, it would be possible to estimate the NH3
losses from effluent-irrigated sites from the effluent's ammoniacal-N together with
the evaporation. The results show that NH3 losses can be important following
effluent-irrigation. Losses of NH3 may be maximised by frequent applications of
a small volume of effluent.
Acknowledgments
This study was supported by funds from the Land and Water Resources
Research and Development Corporation, Murray Darling Basin Commission, New
South Wales Public Works Department, and CSIRO Aust. National Priorities
scheme. We also acknowledge the expert technical assistance provided by Evelyn
A. Colvin.
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Manuscript received 12 February 1996, accepted 30 May 1996