1. No category

Atmospheric Environment 60 (2012) 18e24
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Atmospheric Environment
journal homepage: www.elsevier.com/locate/atmosenv
An experimental determination of the H2S overall mass transfer coefficient from
quiescent surfaces at wastewater treatment plants
Jane Meri Santos a, *, Virginie Kreim b, Jean-Michel Guillot b, Neyval Costa Reis Jr. a, Leandro Melo de Sá c,
Nigel John Horan d
a
Departamento de Engenharia Ambiental, Universidade Federal do Espírito Santo, Av. Fernando Ferrari 514, 29.060-970 Vitória, ES, Brazil
Laboratoire Génie de L’environnement Industriel, École des Mines d’Ales, 6av de Clavières, 30319 Alès Cedex, France
Instituto Federal de Educação, Ciência e Tecnologia do Espírito Santo, Av. Arino Gomes Leal 1700, 29.700-603 Colatina, ES, Brazil
d
School of Civil Engineering, The University of Leeds, Leeds LS2 9JT, UK
b
c
h i g h l i g h t s
< Volatilization of H2S was investigated under different flow conditions.
< Friction velocity does not affect volatilization of H2S at low wind speed.
< WATER9 model provided more realistic estimates of volatilization of H2S.
< Emission models overestimates overall mass transfer coefficient of H2S.
< Volatilization of H2S can be treated as constant for low wind speed conditions.
a r t i c l e i n f o
a b s t r a c t
Article history:
Received 1 July 2011
Received in revised form
5 June 2012
Accepted 7 June 2012
This study has investigated overall mass transfer coefficients of hydrogen sulphide from quiescent liquid
surfaces under simulated laboratory conditions. Wind flow (friction velocity) has been correlated with
the overall mass transfer coefficient (KL) of hydrogen sulphide in the liquid phase using a wind tunnel
study. The experimental values for this coefficient have been compared with predicted KL values
obtained from three different emission models that are widely used to determine volatilization rates
from the quiescent surfaces of wastewater treatment unit processes. Friction velocity (in a range of 0.11
and 0.27 m s1) was found to have a negligible influence on the overall mass transfer coefficients for
hydrogen sulphide but by contrast two of the models predicted a stronger influence of friction velocity
and overestimate the KL values by up to a factor of 12.5, thus risking unnecessary expenditure on odour
control measures. However, at low wind speeds or friction velocities, when more odour complaints
might be expected due to poor atmospheric dispersion, a better agreement of emission rates with
experimental data was found for all the models.
Ó 2012 Elsevier Ltd. All rights reserved.
Keywords:
Mass transfer coefficient
Hydrogen sulphide
Volatilization
Odour
Quiescent surface
1. Introduction
A quiescent surface is characterized by a low level of disturbance
at the aireliquid interface and such surfaces are common at
wastewater treatment plants. In wastewater treatment plants,
primary and secondary settlement tanks are designed to be
quiescent and at larger urban treatment plants these can provide
a large surface area. In addition, certain secondary treatment
* Corresponding author. Tel./fax: þ55 27 33352648.
E-mail
addresses:
[email protected],
[email protected]
(J.M. Santos), [email protected] (V. Kreim), [email protected]
(J.-M. Guillot), [email protected] (N.C. Reis), [email protected] (L.M. de Sá),
[email protected] (N.J. Horan).
1352-2310/$ e see front matter Ó 2012 Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.atmosenv.2012.06.014
options, such as sequencing batch reactors during the settle and
decants phases and biological aerated filters awaiting backwash,
also have large quiescent surfaces. Also, smaller rural works which
employ lagoon-based treatment systems such as waste stabilisation ponds provide much larger quiescent areas. These can all
represent a significant source of atmospheric odour emissions
(Hudson and Ayoko, 2008).
As a result of urbanization and population increases, wastewater
treatment plants that once offered a large buffer zone outside the
site boundary, now find housing developments encroaching within
this buffer zone and often up to the boundary. As a result odour is
an increasingly sensitive issue and complaints of odour nuisance
are common (Nicell, 2009; Latos et al., 2011). Although odourous
emissions during wastewater treatment may comprise a large
J.M. Santos et al. / Atmospheric Environment 60 (2012) 18e24
number of compounds, hydrogen sulphide (H2S) is the most
common source of complaint. It has an odour threshold concentration as low as 0.5 ppb (WEF and ASCE, 1995; Hobson and Yang,
2001), which is often referred to as 1 odour unit (OU) and is
known to cause nuisance at values between 4 and 8 OU (Hobson
and Yang, 2001). As a result H2S is usually employed as a marker
for odour due to its very low detection and recognition thresholds
and its considerable emission rate. In addition, Gostelow and
Parsons (2000) have presented a number of advantages for using
H2S as an indication of the overall odour concentration, such as, it is
easily and rapidly measured down to low ppb levels by hand-held
equipment; gas-phase H2S concentrations can be related to liquid
phase measurements and theoretical models of sulphide formation.
The use of H2S as an odour surrogate in odour modelling is an
important tool during treatment plant design to minimise potential
impacts from odorous unit processes and also as a plant operational
tool, for instance to identify those periods when maintenance tasks
associated with odour release, can be undertaken with minimum
nuisance. In the light of the significant capital expenditures that are
based on the outputs from odour emission and dispersion models,
such models warrant careful attention to ensure their veracity.
Volatilisation describes the process whereby a compound (in
this case the odourant) is transferred from an area source such as
a primary tank surface to the atmosphere. A literature review
indicates that for non aerated quiescent surfaces, volatilization is
usually modelled based on Fick’s law of molecular diffusion and
Henry’s law. The two films theory assumes that a thin layer of liquid
and gas exists where molecular diffusion dominates (Fig. 1). The
mass flux (J) can then be written as a product of a difference
between concentrations and a mass transfer coefficient which
differs for the liquid layer (kL) and the gas layer (kG):
*
*
¼ kG Cair
J ¼ kL Cwater Cwater
Cair
(1)
where C and C* are the concentration at the border of the thin layer
and at the interface between liquid and gas, respectively. However,
as the concentration of odourant at the interface cannot be
measured directly, it is assumed from Henry’s law that
*
*
*
Cair
¼ ðH 0 =RTÞCwater
¼ HCwater
(2)
0
where H is the dimensional form of the Henry’s coefficient ratio
represented as the ratio between the vapour pressure and the
1
concentration in the liquid phase with unit as atm (moll dm3
l ) .R
is the universal gas constant, T is temperature in Kelvin and H is
a dimensionless form of Henry’s coefficient. Then the flux can be
written as
Bulk Gas Phase
C
C*
Interface
Emission (J )
Turbulent Transfer
Molecular
Transfer
Gas Film
C*
Molecular
Transfer
Liquid Film
C
Bulk Liquid Phase
Turbulent Transfer
Fig. 1. . Schematic representation of mass transfer process across liquid and gas films.
C
J ¼ KL Cwater air
H
19
(3)
with
KL ¼
1
1
þ
kL HkG
1
(4)
In this way the overall mass transfer coefficient (KL) becomes
a lumped parameter that incorporates the effects of Henry’s law
together with the individual mass transfers through the liquid and
gas films. It is apparent from Equation (4) that for a given Henry’s
law coefficient the mass transfer is controlled either by liquid phase
(kL) or gas phase (kG). Comparison of Henry’s coefficient has shown
that transfer of polar compounds is controlled by air-phase and
apolar compounds controlled predominantly by liquid-phase. Thus
the magnitude of the mass transfer is a function of the properties of
the odourant being modelled. It is usually modelled empirically
using equations that depend on its physical properties, the
composition of the liquid and gas, the liquid surface geometry and
flow conditions.
However, much of the experimental data that underpin these
empirical equations has been derived from studies using volatile
organic compounds (VOCs) (Trapp and Harland, 1995; Chao et al.,
2005) or ammonia (Bajwa et al., 2006; Rong et al., 2009). VOCs
are low solubility, non-polar compounds with a value for Henry’s
coefficient that indicates emission rate is dependent primarily on
liquid phase (Hudson and Ayoko, 2008). Ammonia is a water
soluble polar compound with a Henry’s constant indicating emission rate is dependent on both air and liquid phase. By contrast
hydrogen sulphide is less polar, sparingly soluble and with an
atypical Henry’s law coefficient (eg: Arogo et al., 1999; Blunden
et al., 2008). Sander (1999) gives a compilation of Henry’s law
0
coefficient (H ) for hydrogen sulphide from 9 references (ranging
between 1.0 103 and 1.0 101 moll dm3
atm1) from which
l
six references give the same value of 1.0 101 moll dm3
atm1
l
5
3
1
(or 3.36 10 Pa m kg ). As such its emission rate is dependent
primarily on liquid phase (Hudson and Ayoko, 2008) and there is
a paucity of experimental data for the mass transfer of this
compound despite it being the most important source of odour
from wastewater treatment plants.
Several emission models have been applied to odour modelling
of sewer networks and wastewater treatment plants, which require
empirical derivations of kL and kG. These include BASTE (Corsi and
Card, 1991), CINCI (Govind et al., 1991), WATER8 (USEPA, 1994),
WATER9 (USEPA, 2001), TOXCHEM (Melcer et al., 1994),
TOXCHEMþ (Enviromega, 2004) and Gostelow et al. (2001). Of
these, perhaps the most widely applied are WATER9, TOXCHEMþ
and the Gostelow model. With all of the models, the overall mass
transfer coefficient has an important role in determining volatilization rates, and as a result, it is important that it is evaluated
accurately. The sub-models that determine the overall mass
transfer coefficient from quiescent surfaces in these three emission
models all follow a different approach. The mass transfer coefficient
for the liquid phase (kL) in quiescent surfaces is calculated in
WATER9 using expressions proposed by Springer et al. (1984) and
Mackay and Yeun (1983) as described in USEPA (1994) that take
account of the fetch-to-depth (F/D) ratio, wind speed (U10) and
molecular diffusivity of the chemical (hydrogen sulphide) in the
water or the Schmidt number for the liquid phase (ScL). These
expressions are generally thought to overestimate the emission
rate (Ferro and Pincince, 1996a and 1996b; Gostelow et al., 2001).
The mass transfer coefficient for the gas phase (kG) is determined as
suggested by Mackay and Matsugu (1973) and it depends on the
Schmidt number of the gas phase (ScG), wind speed and the free
20
J.M. Santos et al. / Atmospheric Environment 60 (2012) 18e24
surface area. The TOXCHEMþ model (Enviromega, 2004) is a more
recent version of the TOXCHEM model proposed by Melcer et al.
(1994). It calculates kL and kG using the expressions proposed by
Mackay and Yeun (1983). As described in Enviromega (2004), these
expressions take into account the friction velocity and ScL or ScG,
but do not use different equations to take account of the F/D ratio.
Gostelow et al. (2001) presented an emission model including
volatilization from quiescent surfaces which considers the liquid
and gas phases mass transfer coefficients as a linear function of
friction velocity and take account of the Schmidt number (ScL or
ScG), according to Mackay and Yeun (1983). The authors estimated
the emission rate of hydrogen sulphide and compared this value to
the ones obtained using WATER8 and TOXCHEMþ models. They
found that for quiescent surfaces such as sedimentation tanks, the
WATER8 model overestimated the emission rate for relatively small
water bodies.
In the light of the increasing importance of H2S volatilization
and in the absence of good quantitative data to describe this, this
study was undertaken to provide additional information on the
factors affecting emission rates of hydrogen sulphide from quiescent liquid surfaces by means of a wind tunnel study. In addition,
this study examines the influence of wind speed (friction velocity)
and other environmental parameters on the H2S overall mass
transfer coefficient. Furthermore this provides the opportunity to
compare the experimentally derived values KL with those predicted
using the equations employed in the three different models
described above.
2. Materials and methods
2.1. Wind tunnel set up
Experiments were conducted using the wind tunnel facility of
the Industrial Environment Engineering Laboratory of the École des
Mines d’Alés (EMA) in France (Fig. 2). The tunnel has a working
section 9.0 m long, 1.0 m wide and 0.5 m high (zone B). The upwind
part (zone A) is used to generate a simulated atmospheric boundary
layer and the experiments are conducted in zone B within the
simulated boundary layer. The wind tunnel is equipped with a blade
fan connected to a frequency variator (Leroy Somer FMV2107) to
produce wind speeds from 0.5 to 4.5 m s1. The main features of this
tunnel are the atmospheric boundary layer simulation in the
approaching flow and the tank (125 cm 60 cm 5 cm) which
simulates the area source with F/D ratio equal to 25. The stainless
surface of the tank was covered with a TedlarÒ film in order to avoid
any reaction of species.
According to Kato and Hanafusa (1996), if the similarity of
vertical distribution of mean wind speed can be obtained, then the
similarity of all other parameters (such as roughness length, friction
velocity, turbulence intensity, turbulent Reynolds number and
Reynolds stress) are also established. In fact, the former three are
functions of friction velocity except the turbulence spectra which is
Fan
Air outlet
made similar by guaranteeing Reynolds number independence.
Thus, the atmospheric boundary layer was simulated by using
a honeycomb and a screen upstream of the test section (zone B) to
even the air flow. The Reynolds number independence was guaranteed by setting this parameter above 3.8 104. The Reynolds
number is defined here based on the experimental values of free
stream velocity and turbulent boundary layer height in the wind
tunnel. A hot-wire anemometer (VT200 model Kimo Instruments,
France) was used to measure the temperature and wind profile
upstream of the tank with an accuracy of 0.1 C and 0.01 m s1. The
tank was filled with 1.0 L of solution containing sodium sulphide
(Na2S) at 5.22 g of Na2S.9H2O/L to produce H2S with an overall
reaction of
Na2 S þ 2H2 O4H2 S þ 2NaOH
(5)
The solution into the tank was diluted with distilled water to
10 mg L1 of H2S and acidified with H2SO4 to pH 4 (measured using
a pH/ion meter accumet model 25, Fisher Scientific) to ensure that
no intermediate products were formed. The use of Na2S diluted
with distilled water to determine the mass transfer coefficient in
the experiment is an adequate representation of wastewater at
treatment plants. Many authors (Stenstrom and Gilbert, 1981;
Chern et al., 2001) indicate that there is a linear correlation
between KL for wastewater and clean water, although the correlation coefficient depends on wastewater quality, suspended solids
concentration, among other factors.
As in a typical experiment with the wind tunnel, temperature
(T air ) and wind profile were measured during the trials. The friction
velocity (U*) and wind speed at the boundary layer height (UN)
were determined assuming a logarithmic wind speed profile
(Table 1). The relative humidity (RH) of air was measured using
a thermo-hygrometer (HD100 model Kimo Instruments, France)
and absolute humidity (AH) of air was estimated from the relative
humidity. The temperature of the liquid (T L ) was measured at
a depth between 2 and 3 cm beneath the surface using a thermohygrometer. Molecular diffusivity of hydrogen sulphide in water
(DL) was calculated according to USEPA (2001) whereas water
viscosity (mL) and water density (rL) were calculated according to
Lide and Kehiaian (1994) (Table 1). The Henry’s coefficient for H2S
was described as a function of liquid temperature according to
expression proposed by Sander (1999).
2.2. Measurements of hydrogen sulphide in the liquid phase
The liquid phase was collected at the interface of the surface
with sample tubing located in three different positions along the
centreline in the wind direction. A drop (0.05 mL) of NaOH solution
was added to each sample to raise the pH within the range 9.5e10
to ensure that only the HS form was prevalent. The conversion of
all residual H2S in the solution to the ionic HS form was then
quantified by UV-spectrometry. HS has a maximum absorbance
at 231 nm (Pouly et al., 1999) and this wavelength is specific to the
Tank simulating area source
Air inlet
0.5 m
Zone C
Zone B
0.75 m 1.25 m
Zone A
7.0 m
Fig. 2. Schematic representation of the wind tunnel used in this study.
1m
J.M. Santos et al. / Atmospheric Environment 60 (2012) 18e24
21
Table 1
Experimental conditions for each of the four friction velocities evaluated.
Trial number
U* (m s1)
UN (m s1)
T L ( C)
T air ( C)
RH (%)
AH ð103 kgH2 O =kgair Þ
DL (109 m2 s1)
mL (103 kg m1 s1)
rL (102 kg m3)
1
2
3
4
5
6
7
8
9
10
11
12
13
14
0.11
0.11
0.11
0.16
0.16
0.16
0.16
0.16
0.16
0.16
0.16
0.21
0.21
0.27
2.30
2.30
2.30
3.45
3.45
3.45
3.45
3.45
3.45
3.45
3.45
4.60
4.60
5.75
17.7
17.5
17.1
18.0
18.8
17.7
17.4
17.3
17.2
17.3
17.5
21.5
19.9
13.7
25.7
21.4
19.1
24.6
18.9
19.2
19.2
19.7
21.1
19.7
22.0
29.8
27.0
17.8
34.0
29.1
39.9
45.4
55.3
44.7
33.4
36.8
19.0
35.8
37.3
38.8
46.5
54.7
6.97
4.60
4.60
8.71
7.52
6.16
4.60
5.22
2.69
4.64
5.69
1.01
1.04
6.88
2.190
2.188
2.185
2.192
2.198
2.190
2.187
2.187
2.186
2.187
2.188
2.218
2.206
2.160
1.065
1.071
1.082
1.057
1.036
1.065
1.073
1.076
1.079
1.076
1.071
0.970
1.008
1.183
9.678
9.679
9.681
9.677
9.674
9.678
9.680
9.680
9.681
9.680
9.679
9.662
9.669
9.695
measured anion, therefore no interference was possible (the
distilled water had no absorbance at 231 nm due to H2SO4 or
NaOH). Calibration of the UV-spectrometer (Antelie model, Secomam, France) was carried out with six solutions of Na2S.9H2O in
a range from 4 to 16 mg L1 of H2S which also had their pH values
increased to between 9.5 and 10 to assure that all sulphur in the
solutions were in the form of HS.
2.3. Estimation of mass transfer coefficient using the liquid phase
measurements
To determine the mass transfer coefficient experimentally, the
estimated flux J was written as suggested by Arogo et al. (1999).
This flux is taken as the temporal variation of the total sulphur (TS)
in the liquid phase. TS ¼ CH2 S , since as previously described, the
liquid phase pH has been modified to ensure that all sulphur is in
the form of H2S, which is important as only H2S volatilizes.
Thus, the estimated flux J is written as
dðTSÞ c
JA ¼ dt
(6)
where A and c are the interfacial area and volume of tank,
respectively. Substituting the Equation (3) into Equation (6) gives
dðTSÞ
A
C
¼ KL
Cwater air :
dt
c
H
(7)
By knowing the pH of the solution one can determine the
partition between sodium sulphide and hydrogen sulphide. Thus,
Cwater ¼ aðTSÞ:
(8)
As the solution used for the experiments was at a pH between 3
and 4, the value of a is equal to unity. Thus, Equation (7) can be
rewritten as
dðTSÞ
A
C
¼ KL
TS air
dt
c
H
(9)
which has the following solution when Cair is neglected:
t
TS ¼ TSt¼0 e
Equation (10). This equation was used to calculate the appropriate
KL values for H2S (Table 2). A multiple linear regression analysis was
conducted to determine which environmental parameters (friction
velocity, temperature difference between the air and liquid, relative
humidity of air, molecular diffusivity of hydrogen sulphide in water,
water viscosity and water density) have a significant effect on the
overall mass transfer coefficient. The results indicate that the
temperature difference between the air and liquid (p ¼ 0.003),
molecular diffusivity of H2S in water (p ¼ 0.033), water viscosity
(p ¼ 0.033) and water density (p ¼ 0.033) were statistically
significant at the 5% significance level, whereas the friction velocity
(p ¼ 0.657) and relative humidity of air (p ¼ 0.077) were not
significant. In addition, the temperature difference between the air
and liquid had a negative effect on the overall mass transfer coefficient, whereas the molecular diffusivity of H2S in water, water
viscosity and water density had a positive effect on the overall mass
transfer coefficient. These results of the statistical analysis indicate
that friction velocity did not have a significant effect on KL values for
H2S. This might be expected for such a sparingly soluble compound
and is also consistent with the Henry’s constant for this compound
which suggests volatilization will depend more on liquid phase
turbulence than wind speed (Schwarzenbach et al., 2003). Arogo
et al. (1999) also noted that KL appeared independent of wind
velocity although Schmidt and Bicudo (2002) considered that
hydrogen sulphide was an exponent of wind speed where the
exponent was between 0.3 and 0.5.
Predicted values of KL were also calculated from the three
models to permit a comparison between experimental and
modelled values (Table 2). For trials 1 to 3, which present smaller
KL A=
c
c
TS
or KL t ¼ ln
A
TSt¼0
(10)
3. Results
The concentration of hydrogen sulphide in the liquid phase was
found to decrease exponentially with time for all the wind velocities tested in this study and this relationship can be described by
Table 2
The influence of wind velocity on the experimentally determined value for KL
compared with the values determined using three common models.
Trial
KL (106 m s1)
number
Water9
Gostelow
Experimental TOXCHEMþ Water9
(Zr ¼ 200 m) (Zr ¼ 1000 m)
1
2
3
4
5
6
7
8
9
10
11
12
13
14
1.28
2.79
2.18
2.03
6.23
5.20
3.66
2.23
4.69
3.15
1.98
2.74
3.86
3.93
5.99
5.98
5.95
12.4
12.5
12.4
12.3
12.3
12.3
12.3
12.3
22.8
22.3
34.9
5.23
5.23
5.23
5.24
5.25
5.23
5.23
5.23
5.23
5.23
5.23
5.28
5.26
5.19
5.23
5.23
5.23
5.24
5.25
5.24
5.23
5.23
5.23
5.23
5.23
5.28
5.26
4.60
17.1
17.0
16.9
24.9
25.2
24.8
24.7
24.7
24.7
24.7
24.8
34.3
33.6
39.6
22
J.M. Santos et al. / Atmospheric Environment 60 (2012) 18e24
values of friction velocity, the models predictions are in better
agreement with experimental data. On the other hand, for trials 12
to 14, which present larger values of friction velocity, the models
predictions were up to 12.5 times larger than those obtained by the
experiments. Whereas the WATER9 model treated KL as a virtual
constant with a value of 5.23 106 m s1 for all except the highest
wind velocity, by contrast both TOXCHEMþ and Gostelow models
returned much higher KL values that increased with increasing
wind velocities (Table 2). This results from the respective treatment
of KL in these models whereby both Gostelow and TOXCHEMþ take
into account the friction velocity to estimate kL for all wind speeds.
By contrast the WATER9 model uses an equation which is not
dependent on friction velocity when the wind speed at 10 m is less
than 3.25 m s1.
The WATER9 model is the only model that uses the values of
velocity at 10 m and thus it is necessary to scale the boundary layer
height (Zr) inside the wind tunnel. This is normally used at one’s
discretion, and so the model was run with two boundary layer
heights (200 m and 1000 m) as described in Table 2. However this
gave no significant change except for when the friction velocity was
equal to 0.27 m s1 which corresponds to a free stream velocity of
5.75 m s1 (Table 1). In this case, the WATER9 employs an equation
which also depends on wind speed (U10), differently from all other
trials.
The low values of KL observed for the WATER9 model in
comparison with other models conflicts with other reported findings. Ferro and Pincince (1996a and 1996b) and Gostelow et al.
(2001) evaluated emission rates of toluene and H2S from
a primary clarifier and sedimentation tank, respectively, using
three different emission models and found that WATER8 predicts
a higher emission rates than the other two models. According to the
authors, WATER8’s calculation of kL depends on mean water
velocity and the depth of the stream and because the water surface
of clarifiers and sedimentation tank are quiescent, this correlation
is inappropriate and overestimates emission rates from these
treatment units with quiescent surfaces. This problem was also
verified by the authors of this present work in the WATER9 code,
which models public clarifiers using correlation for kL indicated for
trenches (channels) (USEPA, 2001). However, in the present work,
neither the WATER8 nor WATER9 software was used, and all the
calculations were performed independently using the actual
equations for quiescent surfaces around which the models are
based. Thus, in this work the wind speed is less than 3.25 m s1 and
WATER9 model estimates kL using an equation which is not
dependent on wind velocity. As a result estimates for KL are
approximately constant.
Fig. 3 shows the effect of the friction velocity on the experimental and modelled overall mass transfer coefficients (KL) for
hydrogen sulphide. The differences between Gostelow, TOXCHEMþ
and WATER9 predictions shown in Fig. 3 can be explained from the
approaches used to calculate kL since the volatilization of hydrogen
sulphide is dominated by mass transfer in the liquid phase. The
Gostelow model overestimates H2S mass transfer by an order of
magnitude in all experiments. This is because this model employs
a linear function of friction velocity to calculate liquid phase mass
transfer coefficient at all wind speed. This equation was obtained by
Mackay and Yeun (1983) for friction velocity greater than 0.3 m s1
and overestimates kL when used at lower wind flow conditions,
such as in this present study. Unlike Gostelow, TOXCHEMþ calculates kL using a power function of friction velocity with exponent
equal to 2.2 (indicated for U* < 0.3 m s1), which is appropriate for
the experimental conditions employed in this work. Thus, the
TOXCHEMþ model presents an overall mass transfer coefficient in
better agreement with the experimental data than that of Gostelow, especially for the trials with lower friction velocity
(U* ¼ 0.11 m s1). The mass transfer coefficients calculated using
the WATER9 model was in reasonable agreement with experimental results. This agreement may be due to two factors: (1) the
WATER9 model calculates the liquid phase mass transfer coefficient
for low wind speed (U10 < 3.25 m s1) using an equation which is
not dependent on wind speed (depends on the ratio between
molecular diffusivity of hydrogen sulphide and ether in the water);
and (2) the friction velocity did not have a significant effect on
experimental KL, as found by Arogo et al. (1999).
It is important to note that the formulations employed by the
models to calculate kL and kG were obtained under different
experimental conditions to this work, except for Henry’s constant
which is within the range investigated by Mackay and Yeun. In this
work, the friction velocity varied between 0.11 and 0.27 m s1 and
Schmidt number in the liquid film (ScL) was between 452 and 564
whereas in Mackay and Yeun’s work the friction velocity varied
between 0.27 and 0.9 m s1 and the Schmidt number in the liquid
film was between 939 and 1340. In other words, the values of U* in
this work were 3e8 times smaller than those in the Mackay and
45
40
35
25
-6
KL (10 m/s)
30
20
15
Experimental (this study)
WAT ER9 (Zr=200 m)
WAT ER9 (Zr=1000 m)
T OXCHEM+
Gostelow et al. (2001)
10
5
0
0
0.05
0.1
0.15
0.2
0.25
U* (m/s)
Fig. 3. Effect of friction velocity on the overall mass transfer coefficient of hydrogen sulphide.
0.3
J.M. Santos et al. / Atmospheric Environment 60 (2012) 18e24
Yeun’s experiments and ScL were 2e2.5 times smaller. Thus, for the
flow conditions and compounds investigated by Mackay and Yeun,
the wind friction velocity may have had a strong influence on
volatilization, whereas for hydrogen sulphide at a lower friction
velocity, this latter parameter has shown no significant influence
on volatilization rates. Furthermore, in the experimental
conditions of this study (U10 < 3.25 m s1), both TOXCHEMþ and
Gostelow models employed Mackay and Yeun’s (1983) expressions to calculate kL and kG. On the other hand, the WATER9 model
uses Springer et al. (1984) and Mackay and Matsugu (1973) equations to calculate kL and kG, respectively, which were obtained from
the experimental conditions similar that the investigated in this
work.
Fig. 4 shows the comparison between the experimental and
modelled overall mass transfer coefficients for hydrogen sulphide.
The strong influence of the friction velocity on the overall mass
transfer coefficient of hydrogen sulphide estimated by the
TOXCHEMþ and Gostelow was responsible for the grouping points
in Fig. 4. For each of these two models, four separate clusters of
points corresponding to four different values of friction velocity
appear (Fig. 4). The number of points in each cluster is equivalent to
the number of experiments conducted under the same friction
velocity. The WATER9 model gave a single cluster of points due to
its independence on wind speed when calculating the coefficient of
mass transfer in the liquid phase for the range of friction velocity
investigated. The grouped points (from trials with the same friction
velocity) are presented in Fig. 4 in a straight line which may indicate that the liquid and air temperature have a weaker influence on
the KL values calculated by the models in comparison to those
obtained in this experimental study, i.e., the influence of temperature may be stronger than that incorporated by the models though
the gas and liquid physical properties. Furthermore, liquid
temperature might have a stronger influence than that foreseen by
the models since volatilization of hydrogen sulphide is dominated
by the mass transfer in liquid phase. The modelled liquid phase
mass transfer coefficient was about 102e103 times smaller than
that gas phase mass transfer coefficient, depending on the model.
Thus, the results of this work indicated that the volatilization of
23
hydrogen sulphide at low wind speed can be considered
independent of wind velocity, and that the TOXCHEMþ and
Gostelow models give considerable overestimation for mass
transfer coefficient of H2S.
4. Conclusions
The important factors that influence the overall mass transfer
coefficient for the volatilization of hydrogen sulphide from quiescent surfaces have been quantified experimentally. Existing emission models, originally developed for volatile organic compounds
and ammonia, have been assessed for hydrogen sulphide and the
WATER9 model provided more realistic estimates of the mass
transfer coefficient when compared to other models. However,
a new equation to calculate KL for hydrogen sulphide should be
developed though experiments and dimensional analysis and
included in these existing models.
Friction velocity does not have a significant effect on the overall
mass transfer coefficients for hydrogen sulphide in the wind tunnel
experimental results and so it can be ignored, in contrast to results
from both the TOXCHEMþ and Gostelow et al. models which
overestimate the size of the overall mass transfer coefficient due to
the strong influence of friction velocity in their models.
WATER9 showed the best agreement with experimental data
although it was found to overestimate the overall mass transfer
coefficient by a factor of up to 4.0 and TOXCHEMþ only gave good
agreement at low wind speed or friction velocity.
The influence of the liquid and air temperature on the experimental KL values obtained in this study may be stronger than that
incorporated by the models though the gas and liquid physical
properties.
Volatilization of hydrogen sulphide can be treated as approximately constant in low wind speed, i.e. friction velocity less than
0.3 m s1, according to WATER9. For higher wind speeds, H2S may
exhibit volatilization dependent on wind speed as suggests the
volatilization models.
Acknowledgements
The authors wish to acknowledge the sponsorship of the Brazilian Government through CAPES (Fundação Coordenação de
Aperfeiçoamento de Pessoal de Nível Superior).
40.0
References
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Calculated KL (10 m/s)
30.0
20.0
10.0
T OXCHEM+
WAT ER9 (Zr=200 m)
WAT ER9 (Zr=1000 m)
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Experimental K L (10 m/s)
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