Journal of Environmental Sciences
Volume 30 2015
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J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 5 5– 6 4
Available online at www.sciencedirect.com
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www.journals.elsevier.com/journal-of-environmental-sciences
Abatement of SO2–NOx binary gas mixtures using a
ferruginous active absorbent: Part I. Synergistic effects
and mechanism
Yinghui Han1,2,3 , Xiaolei Li 4 , Maohong Fan 2,3,⁎, Armistead G. Russell 3 , Yi Zhao1 ,
Chunmei Cao1 , Ning Zhang1 , Genshan Jiang 1
1.
2.
3.
4.
School of Mathematics and Physics, North China Electric Power University, Baoding 071003, China
School of Energy Resources and Department of Chemical & Petroleum Engineering, University of Wyoming, Laramie, WY 82071, USA
School of Civil and Environmental Engineering, Georgia Institute of Technology, Atlanta, GA 30332, USA
Shanxi Electric Power Exploration & Design Institute, China's State Energy Group Co., Taiyuan 030001, China
AR TIC LE I N FO
ABS TR ACT
Article history:
A novel ferruginous active absorbent, prepared by fly ash, industrial lime and the additive
Received 20 May 2014
Fe(VI), was introduced for synchronous abatement of binary mixtures of SO2–NOx from
Revised 29 September 2014
simulated coal-fired flue gas. The synergistic action of various factors on the absorption of
Accepted 11 October 2014
SO2 and NOx was investigated. The results show that a strong synergistic effect exists
Available online 21 February 2015
between Fe(VI) dose and reaction temperature for the desulfurization. It was observed that
in the denitration process, the synergy of Fe(VI) dose and Ca/(S + N) had the most
Keywords:
significant impact on the removal of NO, followed by the synergy of Fe(VI) and reaction
Fe(VI)
temperature, and then the synergy of reaction temperature and flue gas humidity. A
Desulfurization
scanning electron microscope (SEM) and an accessory X-ray energy spectrometer (EDS)
Denitration
were used to observe the surface characteristics of the raw and spent absorbent as well as
Binary gas
fly ash. A reaction mechanism was proposed based on chemical analysis of sulfur and
Synergized effect
nitrogen species concentrations in the spent absorbent. The Gibbs free energy, equilibrium
Thermodynamics
constants and partial pressures of the SO2–NOx binary system were determined by
thermodynamics.
© 2015 The Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences.
Published by Elsevier B.V.
Introduction
Emissions of sulfur dioxide (SO2) and nitrogen oxides (NOx) from
the electric power sector are considered two of the major causes
for the deterioration of the atmosphere environment in China
(Hu et al., 2010; Lin et al., 2010; Lu et al., 2010). In an effort to abate
SO2 and NOx emissions, a number of coal-fired retrofitting units
have been investigated, such as flue gas desulfurization (FGD)
and wet-scrubbers. There has also been an effort by coal-fired
plants to switch to coal of lower sulfur content. These measures
alleviate some of the pressure on SO2 air pollution, but the
concentration of NOx in the atmosphere has not decreased. In
2011, the total emission of NOx in China was 1206.7 million tons,
jumping 6.17% from a year earlier (Han, 2012). Even worse, the
area of acid rain never decreased; the main kind of acid rain was
converted from simple sulfuric acid to a complex type of sulfuric
acid and nitrating acid (Hu et al., 2010). In recent years, some
effort has been expended to limit NOx emissions, i.e., selective
catalytic reduction (SCR), selective non-catalytic reduction
(SNCR), low NOx burners, and so on; however, such denitration
⁎ Corresponding author. E-mail: mfan@uwyo.edu (Maohong Fan).
http://dx.doi.org/10.1016/j.jes.2014.10.012
1001-0742/© 2015 The Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences. Published by Elsevier B.V.
56
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 55–6 4
equipment installed at the front or the back of the FGD system
has some disadvantages such as large footprint and high
running cost (Zhao et al., 2012).
A great deal of research work aimed at synchronizing
abatement of SO2 and NOx has been carried out recently (Xie et
al., 2004; Yi et al., 2012; Yuan et al., 2012; Ding and Yazaydin, 2013;
Krishnan et al., 2013), but many technical and economic
problems have been encountered, impairing the implementation
of practicable technologies. Therefore, it is still significant
and urgent to develop high efficiency, practical techniques of
synchronizing desulfurization and denitration, without introducing secondary pollution, according to the characteristics of
coal-fired flue gas, especially those compatible with the existing
FGD systems. The linchpin of synchronizing abatement of SO2
and NOx techniques is to oxidize NO into NO2 rapidly in the flue
gas, as the latter is readily dissolved in water. There have been
several studies on the rapid oxidation of NO (Deshwal et al., 2008;
Li et al., 2011; Liu and Zhang, 2011), however, some strong
oxidizing absorbents were not only economically infeasible, but
also tended to introduce heavy metal secondary pollution during
industrial application. In order to solve these practical problems,
a new eco-friendly oxidizing additive, Fe(VI), was introduced into
a flue gas circulating fluidized bed system in a previous study by
the authors (Zhao et al., 2011). Fe(VI) is one of the most promising
environmentally benign chemicals and will not produce any
harmful byproducts (Sharma, 2010). It has strong oxidizing
power and its redox potential ranges from 2.20 V in an acidic
environment to 0.72 V in a basic environment. Such outstanding
oxidizing performance and environmentally benign nature
make Fe(VI) a preferred choice in the pollutant disposal process.
The addition of green oxidizing Fe(VI) was effective in improving
the synchronization of pollutant removal abilities in flue gas
control and purification, which has also recently been confirmed
by many researchers (Sharma, 2010; Xia et al., 2011; Dutcher et
al., 2011). However, these investigations were limited. As is well
known, there are many operational factors that will affect the
removal of SO2 and NO in real power plants. In conventional
multi-factor experiments, optimization is usually performed by
varying a single factor while keeping all other factors fixed under
a specific set of conditions. This is not only time-consuming, but
also usually incapable of reaching the true optimum due to
neglect of the synergistic effects among factors. These shortcomings might restrict the practical application of the promising
Fe(VI)-based removal technology. To fill this gap, multi-factor
experiments of removing SO2 and NOx were carried out using
the developed ferruginous active absorbent, and the synergistic
action of various factors was investigated in this work.
For the first time, a novel statistical experimental design
method, response surface methodology (RSM), was adopted in
this work. This is because RSM has many inherent advantages
over other experimental design methods, especially in revealing
the synergistic effects between multiple variables with fewer
experimental trials, optimizing technological operation and
guiding industrial management. In addition, the related removal
mechanism still remains enigmatic. In this work, the reaction
mechanisms were revealed based on chemical analysis and
testing techniques such as X-ray diffraction (XRD), scanning
electron microscope (SEM) and Energy dispersive spectroscopy
(EDS). These results will be very useful for improving and
optimizing the flue gas treatment process using Fe(VI) in
practical applications.
1. Experimental
1.1. Preparation and characterization of a ferruginous active
absorbent
In order to meet the demands of industrial applications, the
preparation process strove for simple and quick operation,
without involving stirring, digesting or drying. The ferruginous
active absorbent was prepared by adding an amount of K2FeO4
to a mixture of fly ash and industrial lime with the ratio of 3:1
by weight. The mass percent of the additive, namely, dose of
Fe(VI), ranged from 0% to 1.0%. In addition, tiny amounts of
water were injected into solid mixtures with the solid/liquid
ratio of 10:1 by weight. The fly ash samples were collected from
the electrostatic precipitators of the Matou coal-fired plants in
Hebei, China; the industrial lime was technical-pure grade
(Tianjin Tanggu Chemical Reagent Factory, Tianjin, China); all
other reagents were analytical grade, used without further
purification and purchased from Beijing Chemical Reagent Co.,
China. The compositions of fly ash and industrial lime used in
the work are summarized in Table 1. The specific surface areas
of fly ash, industrial lime, and fresh absorbent were measured
(SA-3100 surface area & pore size analyzer, Beckman Coulter
Inc., USA) by the BET standard method; the results are as follows:
fly ash 4.588 m2/g; industrial lime 10.173 m2/g; and 45.349 m2/g.
The additive, K2FeO4, was characterized by an X-ray diffractometer (D8 Advance type, Bruker Corporation, Germany). An
scanning electron microscope (SEM, KYKY-2800B type, KYKY
Technology Co., LTD., Beijing, China) and an accessory
energy-dispersive X-ray spectroscopy (EDS, NORAN System 7
type, Thermo Fisher Scientific Inc., MA, USA) were used to
observe the surface properties and to determine relative surface
contents of significant elements in the fly ash, lime absorbent, ferruginous highly active absorbent, and its removal
products.
1.2. Apparatus and procedure
The experiments on synchronous abatement of SO2–NOx binary
gas mixtures were performed in a fixed bed reactor, as shown in
Fig. 1, with a vertical cylinder with a length of 250 mm and an
inner diameter of 20 mm. The fixed bed reactor filled with
Table 1 – Composition of fly ash and industrial lime.
Components content (wt.%)
SiO2
Al2O3
Fe2O3
CaO
MgO
Ig-loss
Others
Fly ash
Industrial lime
46.6
0.4
17.2
0.15
11.1
0.11
7.3
72.3
6.4
1.3
10
25.2
1.4
0
57
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 5 5–6 4
Check valve Three-way valve
Gaseous products
Glass
rotameter
Glass
rotameter
Regulator
Mixed buffer bottle
Gas drying bottle
Flue gas analyzer
Waste gas
scrubber
Fixed bed
Furnace
By-pass
Scanning electron
microscope
Controlling valve
SO2
NO
N2
Solid products
Energy-dispersive
X-ray spectroscopy
Fig. 1 – Schematic diagram of the experimental procedure for synchronous abatement of binary SO2–NO.
ferruginous active absorbent was heated by an outer furnace.
The simulated flue gases, SO2 and NO, were supplied from
compressed gas steel cylinders. The flow rates of gases were
adjusted by pressure reducing valves and glass rotameters. The
carrier gas, N2, makes the simulated flue gas well-distributed in
the mixed buffer bottle. The simulated flue gas was then passed
through the system bypass to the absorption bottles. After the gas
steam was steady, the simulated flue gas was introduced to the
fixed-bed reactor by turning the three-way valve. The tail gas was
discharged to the atmosphere after it was absorbed in absorption
bottles, and the three-way valve was then turned again.
The removal efficiencies were calculated from the measured
concentrations of SO2 and NO at the inlet and outlet as
measured by a flue gas analyzer, vario plus industrial type, MRU
Intruments Inc., Germany. The flue gas moisture was measured
by the flue gas analyzer. The reaction temperature was
controlled and measured with a thermocouple built into the
furnace. The solid absorbent and solid products were characterized by an X-ray diffractometer, a SEM and an accessory EDS.
Concentrations of sulfur (sulfate and sulfite) and nitrogen (nitrate
and nitrite) compounds were determined using barium chromate
photometry, and N-(1-naphthyl)-ethylenediamine photometry,
respectively.
1.3. Experimental design and calculation method
To evaluate the degree of significance of various factors and
their synergistic effects, and therefore, to determine the optimal
operating conditions for the removal of SO2 and NO by the
ferruginous active absorbent, a central composite design (CCD)
with 4 factors and 5 levels was employed using response surface
methodology (RSM), which required 30 runs. The orthogonal
factorial design consisted of 16 orthogonal points, 8 star (or axial)
points and 6 center points. The variables and their levels
selected for the removal of SO2 and NO were as follows:
Ca/(S + N) ranged from 0.8 to 1.6; Fe(VI) dose varied from 0.4%
to 1.2% (wt); reaction temperature ranged from 40°C to 80°C;
and flue gas humidity varied from 2% to 10% (The detailed
experimental design and test results are listed in Table S1 and
Table S2 in the Additional Information, respectively.). The other
related experimental reaction conditions were: the flow rate of
the simulated flue gas was 0.12 m3/hr; the packed mass of the
absorbent was 10 g. The inlet concentrations of SO2 and NO
varied from 1600 to 2000 mg/m3 and from 400 to 800 mg/m3,
respectively. All the experimental results were evaluated with
the aid of statistical graphics Design-Expert 8.0.5b software
(Version 8, Stat-Ease, Inc., Minneapolis, MN, USA) in the data
analysis.
The removal efficiencies were calculated by
η¼
C i;in −C i;out
100%
C i;in
ðE1Þ
where, Ci,in represents the concentration of NO or SO2 at the
inlet of the reactor and Ci,out represents the concentration of
NO or SO2 at the outlet of the reactor.
An important factor, Ca/(S + N), as one of the important
indexes effecting the removal efficiencies of SO2 and NO, is
usually used to evaluate the utilization of absorbents and is
expressed as shown below,
m x 90%=74
Ca=ðS þ NÞ ¼ C SO2 =64 þ 1=2ðC NO =30Þ Q 0:001 t
ðE2Þ
where, m (g) represents the loading weight of the absorbent, x
(wt.%) stands for the weight percentage of industrial lime, C SO2
and CNO (mg/m3) are the concentrations of SO2 and NO,
respectively, Q (m3/hr) represents the flow rate, t (hr) is the
reaction time, the other parameters, “90%”, “74”, “64”, and “30”
are the purity of the industrial lime, the molar mass of
Ca(OH)2, the molar mass of SO2 and the molar mass of NO,
respectively. Correspondingly, the reciprocal of Ca/(S + N) can
be estimated as the absorbent utilization (Yang, 1978).
2. Results and discussion
2.1. Synergistic effects of operating factors
The absorption behavior of the Fe(VI) absorbent was investigated
in the previous studies (Zhao et al., 2011; Zhang et al., 2012) and
the independent influences of various factors on the desulfurization and denitration were analyzed. However, single factor
58
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 55–6 4
analysis cannot fully reflect the inherent law and synergistic
action of operating factors on the removal process. As stated
above, the experiments of synchronous abatement of SO2–
NOx binary gas mixtures were designed by response surface
methodology (RSM) to confirm the synergistic effects of four
operating factors (Fe(VI) dose, Ca/(S + N), reaction temperature,
and flue gas humidity) on desulfurization and denitration
efficiencies.
The analysis of experimental results indicated that a strong
synergistic effect exists between Fe(VI) dose and reaction
temperature for the desulfurization process. In the denitration
process, the synergy of Fe(VI) dose and Ca/(S + N) had the most
significant impact on the removal of NO, followed by the synergy
of Fe(VI) and reaction temperature, and then the synergy of
reaction temperature and flue gas humidity. These synergistic
effects are not a simple addition of various single factors, but a
complex nonlinear relationship. Fig. 2 shows the significant
synergistic effects. It can be noticed that both the desulfurization
efficiency and the denitration efficiency increased rapidly with
the Fe(VI) dose, as shown in Fig. 2a, b and c, which indicate
that the Fe(VI) additive supplies the “oxidizing-points” on the
a
absorbent surface. SO2 and NO from flue gas can be removed and
converted into high valence state sulfur and nitrogen by these
“oxidizing-points”.
Fig. 2a, c shows that reaction temperature acts synergistically with Fe(VI) dose for both desulfurization and denitration.
Synchronous desulfurization and denitration is a semidry
reaction process between porous solids and gases. Reaction
temperature has a great effect on adsorption, absorption and
diffusion of SO2 and NO on the surface of the ferruginous active
absorbent. The optimal temperature was observed to be 60°C in
this experiment. The existence of an optimal temperature can
be explained by the following: when the temperature is lower
than a critical value, adsorption and absorption are predominant and the efficiencies increase with the rise of reaction
temperature. When the reaction temperature is higher than a
critical value, desorption of gas molecules on the surface of the
absorbent increases and the equilibrium adsorptive capacity of
gas decreases. The growth trend of removal efficiencies slows
with the rise of temperature, particularly for denitration. This
critical value of the reaction temperature could be the optimum,
and it will also affect the oxidizing properties of Fe(VI). A
b
)
VI
(
Fe
e
dos
t)
,w
(%
Fe
(V
I) d
ose
(%
Ca/(S+N)
,w
t)
d
c
Fe
(
ose
)d
VI
t)
,w
(%
Flue
gas
hu
mid
ity (
%)
em
nt
tio
ac
Re
)
°C
e(
tur
ra
pe
Fig. 2 – Synergized effects on the desulfurization and denitration, respectively. (a) Synergy of Fe(VI) dose and reaction temperature vs.
desulfurization efficiencies, in which the flow rate was 0.12 m3/hr; the concentration of SO2 was 1600 mg/m3; the flue gas humidity
was 6%; and Ca/(S + N) was 1.1, respectively; (b) synergy of Fe(VI) dose and Ca/(S + N) vs. denitration efficiencies, in which the flow
rate was 0.12 m3/hr; the concentration of NO was 700 mg/m3; the flue gas humidity was 6%; and the reaction temperature was 60°C,
respectively; (c) synergy of Fe(VI) dose and reaction temperature vs. denitration efficiencies, in which the flow rate was 0.12 m3/hr;
the concentration of NO was 700 mg/m3; the flue gas humidity was 6%; and Ca/(S + N) was 1.1, respectively; (d) synergy of reaction
temperature and flue gas humidity vs. denitration efficiencies in which the flow rate was 0.12 m3/hr; the concentration of NO was
700 mg/m3; the reaction temperature was 60°C; and Ca/(S + N) was 1.1, respectively.
59
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 5 5–6 4
2.2. Structural characteristics and product analysis
The SEM analyses of samples (Fig. 4) show that fly ash is
relatively smooth because of its small particle size of tens of
1.0
SO2
NO
0.8
Cout/Cin
temperature that is too high will destroy the stability of Fe(VI);
conversely, a temperature that is too low is detrimental to the
removal reaction. A relatively ideal reaction temperature, 60°C,
was suggested for future industrial application.
Fig. 2a, b, c shows the synergistic effect of Fe(VI) dose together
with reaction temperature and Ca/(S + N) for desulfurization and
denitration. All peaks of the efficiencies appeared at Fe(VI) dose
0.8% (wt). Generally speaking, the more oxidizing additive is
added, the higher the removal efficiencies obtained, especially
when the dose of the oxidizing additive is small. However, when
the dose of the oxidizing additive gets larger, the efficiency is
enhanced more slowly, which can be explained by the saturation
of the “oxidizing points” on the active absorbent (Zhao et al., 2011)
when the dose gets larger than an optimal value. Considering the
synergy of the various effects, the optimal dose of Fe (VI) was
determined as 0.8% (wt.) in this work.
Generally, Ca/(S + N) is often considered as the main factor
for desulfurization, but subsequent analysis of experimental
results revealed that the importance of Ca(S + N) is reflected
only in the mono-factorial influence on desulfurization. It is
seen that Ca/(S + N) has no significant synergistic relationship
with the other factors for desulfurization in this work.
However, unlike desulfurization, the interaction of Ca/(S + N)
and Fe (VI) dose affects denitration to some extent, as shown
in Fig. 2b. The data suggest that oxidation of NO is highly
dependent on the synergistic effect of Fe (VI) additive and
Ca/(S + N), whereby the oxidative product unites with a Ca-base
group to form a calcium salt, resulting in the denitration
efficiency being influenced by both Fe(VI) dose and Ca/(S + N);
however, SO2 reacts with lime easily, even in the absence
of Fe(VI), so there is no significant synergy between Ca(S + N)
and Fe(VI) for desulfurization. The optimal ratio of Ca/(S + N) is
determined as 1.1; correspondingly, the effective utilization of
the absorbent might be estimated as 90.9%.
In addition to reaction temperature, the flue gas humidity
will affect the removal efficiencies, though to a lesser degree.
Fig. 2d shows that the denitration efficiencies increased slightly
with the flue gas humidity, especially at the lower reaction
temperature; but as the flue gas humidity increased, the degree
of increase of the removal efficiencies declined little by little.
Due to the co-constraint of flue gas humidity and reaction
temperature, the 3D surface related to their synergistic effects
on the denitration efficiency was relatively flat.
Under the obtained optimal experimental conditions, i.e., the
dose of Fe(VI) 0.8% (wt), ratio of Ca/(S + N) 1.1, reaction
temperature of 60°C, further parallel experiments were performed, and the time-dependence of the removal rates (Cout/Cin)
of SO2–NO binary gas mixtures was obtained, as shown in Fig. 3.
The results showed the reaction between Fe(VI) and the
simulated flue gas occurred very rapidly. The outlet concentrations of SO2 and NO decreased sharply at the beginning of the
removal process, and then reached plateaus successively after
nearly 10 min; thereafter, the removal rates (Cout/Cin) remained
constant, which indicates the stable removal capability of SO2
and NO by Fe(VI).
0.6
0.4
0.2
0
0
15
30
45
60
75
90
Time (min)
Fig. 3 – Variation curves of the removal rates of SO2–NO binary
gas mixtures with the reaction time under the initial
concentrations of SO2 and NO 2000 and 700 mg/m3, respectively.
Reaction conditions: the flow rate was 0.12 m3/hr; reaction
temperature maintained 60°C, the dose of Fe(VI) was 0.8% (wt),
the ratio of Ca/(S + N) was 1.1.
microns. The other four solids are much coarser, indicating
larger particle size. The pore structure of the fresh absorbent
was observed to be clearer than that of fly ash and industrial
lime particle. From Table 1, it can be seen that Si is the main
component of fly ash. Due to the pozzolanic reaction (Fan and
Brown, 2001), Si compounds in fly ash (Fan et al., 2004) might
react with lime and the surface of fly ash was corroded by
Ca(OH)2, which is the possible reason that the porous structure
of the absorbent appeared. The surface of the spent absorbent is
evidently different from that of the fresh one, and the white
flake layer covered by porous substances might consist of the
removal products, such as calcium sulfate, calcium sulfite,
calcium nitrate and calcium nitrite.
The averaged energy spectrums of the surface of the fresh
ferruginous active absorbent and the spent absorbent are
shown in Fig. 5, respectively. The rising levels of sulfur
content and the appearance of nitrogen content in the spent
absorbent demonstrate that chemical reaction took place in
the flue gas and ferruginous active absorbent. In order to
ascertain the components in the chemical reaction products,
the contents of sulfur and nitrogen species in the raw
absorbent and the spent absorbent were analyzed by a
chemical analysis method, as presented in Table 2. In the
raw absorbent, only trace amounts of sulfate were detected,
probably from fly ash and industrial lime. It is important to
note that no sulfite, nitrate, or nitrite was detected in the raw
absorbent, illustrating that the sulfur and nitrogen components in the spent absorbent were derived from the reaction,
not from the raw absorbent. The sulfate product came
predominantly from desulfurization. The molar ratio of
nitrate/nitrite was found to be 2.06, indicating that there are
two components in the denitration products, with the nitrite
as the main product. This is similar to results obtained in the
previous research in a flue gas circulating fluidized bed (Zhao
et al., 2011).
60
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 55–6 4
a
b
c
d
Fig. 4 – Scanning electron microscope images of samples. (a) Fly ash particle; (b) industrial lime particle; (c) the spent absorbent;
(d) the fresh ferruginous active absorbent, respectively.
2.3. Reaction mechanism
The reaction between SO2–NOx binary gas mixtures and the
ferruginous active absorbent in a fixed bed reactor may be a
complicated physical adsorption and chemical absorption
process, involving many reactions and steps.
In terms of desulfurization, the presence of lime makes the
removal process of SO2 occur quite easily,
CaðOHÞ2 þ SO2 →CaSO3 þ H2 O:
ð1Þ
Here calcium sulfite is regarded as an unexpected product,
because it would diminish the quality of gypsum. The production
of calcium sulfite is the result of insufficient oxidation. According
to an early FT-IR analysis (Li et al., 2002), NO and SO2 can react
with each other, as shown below:
efficiency of NO was zero using the fly ash or industrial lime
alone. The adsorbed NO on the surface of absorbent might be
desorbed or replaced by SO2 compounds, likely because calcium
compounds interact more strongly with SO2 than NOx. Even if NO
could be adsorbed, it is difficult to convert into higher valence
nitrogen. Therefore, the involvement of an oxidant is necessary
for denitration.
In this work, Fe(VI) as an oxidizing additive was dispersed on
the surface of the absorbent, which generated many oxidizing
sites. These oxidizing sites convert SO2 and NO into SO3 and
NO2 rapidly. The fly ash in the absorbent might play a catalytic
role in NO oxidation.
fly ash
NO þ FeðVIÞ → NO2 þ FeðIIIÞ
fly ash
ð4Þ
SO2 þ 2NO→N2 O þ SO3
ð2Þ
SO2 þ FeðVIÞ → SO3 þ FeðIIIÞ:
2SO3 þ NO→ðSO3 Þ2 NO:
ð3Þ
The formed NO2 can react with SO2 or undergo a chain
reaction with itself (Li et al., 2002).
However, the reaction rate of Reaction (2) would be too slow to
produce substantial amounts of high valence sulfur (Armitage
and Cullis, 1971). In the presence of lime, Reaction (1) is more
likely to occur than Reaction (2). Thus, a strong oxidant is needed
to promote the thorough removal of SO2, but this is not the case
for denitration. It was found in this experiment that the removal
ð5Þ
NO2 þ 2SO2 →ðSO3 Þ2 NO
ð6Þ
2NO2 →N2 O4
ð7Þ
N2 O4 þ 2NO→2N2 O3
ð8Þ
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 5 5–6 4
61
keV
Fig. 5 – Energy dispersive spectrums of the ferruginous active absorbent (a) and the spent absorbent (b).
These gaseous products could be further adsorbed physically
on the surface of the ferruginous active absorbent. When the
water vapor in the flue gas was diffused into the surface of
absorbent, NO2, N2O4, SO2, and SO3 were dissolved into the water
film,
3NO2 þ H2 O→2HNO3 þ NO
ð9Þ
NO2 þ NO þ H2 O→2HNO2
ð10Þ
N2 O4 þ H2 O→HNO3 þ HNO2
ð11Þ
HNO2 þ FeðVIÞ→HNO3 þ FeðIIIÞ
ð12Þ
SO2 þ H2 O→H2 SO3
ð13Þ
H2 SO3 þ FeðVIÞ→H2 SO4 þ FeðIIIÞ
ð14Þ
SO3 þ H2 O→H2 SO4 :
ð15Þ
CaðOHÞ2 þ H2 SO4 →CaSO4 þ 2H2 O
ð17Þ
CaðOHÞ2 þ 2HNO2 →CaðNO2 Þ2 þ 2H2 O
ð18Þ
CaðOHÞ2 þ 2HNO3 →CaðNO3 Þ2 þ 2H2 O
ð19Þ
CaSO3 þ FeðVIÞ þ CaðNO2 Þ2 →CaSO4 þ CaðNO3 Þ2 þ FeðIIIÞ:
ð20Þ
The above EDS and chemical analyses demonstrate that the
main products of the desulfurization and denitration are
calcium sulfate and calcium nitrate, with trace amounts of
calcium sulfite and calcium nitrite. On that evidence, we find
that Reactions (5), (14) and (17) are the key steps of desulfurization, while Reactions (4), (10), (11) and (19) are the main
processes of denitration. In addition, according to the analysis
by XRD in the previous work (Zhao et al., 2011), the reduction
product of the Fe(VI) additive is calcium ferrite.
Therefore, the overall reaction can be expressed as below,
fly ash
3SO2 þ 2K2 FeO4 þ 3CaðOHÞ2 → CaSO4 þ Ca2 Fe2 O5 þ 2K2 SO4 þ 3H2 O:
ð21Þ
These higher valence sulfur and nitrogen oxides react with
Ca(OH)2 in the absorbent and form calcium salts,
CaðOHÞ2 þ H2 SO3 →CaSO3 þ 2H2 O
ð16Þ
Under the condition of lack of oxygen,
fly ash
6NO þ 2K2 FeO4 þ 3CaðOHÞ2 → CaðNO2 Þ2 þ Ca2 Fe2 O5 þ 4KNO2 þ 3H2 O
ð22Þ
In an oxygen-enriched environment,
fly ash
6NO þ 2K2 FeO4 þ 3CaðOHÞ2 þ 3O2 → CaðNO3 Þ2 þ Ca2 Fe2 O5 þ 4KNO3 þ 3H2 O:
Table 2 – Contents of sulfur and nitrogen species in the
fresh absorbent and the spent absorbent.
Fresh absorbent (mmol/g)
Spent absorbent (mmol/g)
[SO2−
4 ]
[SO2−
3 ]
[NO−2]
[NO−3]
0.001
0.986
0
0.014
0
0.289
0
0.596
ð23Þ
But in practice, there is 6%–8% oxygen in the flue gas, so the
above denitration processes, Reactions (22) and (23), might occur
concurrently. The product analysis in Table 2 shows that the
denitration products are both calcium nitrate and calcium nitrite,
62
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 55–6 4
but mainly in the form of calcium nitrate. Accordingly, Reaction (23) was inferred as the main reaction.
2.4. Mass balance
The fresh and spent samples of ferruginous active absorbent
were quantitatively analyzed to examine how much of the
absorbent was un-reacted. The formation of sulfur and nitrogen
species as inferred above in Reactions (21)–(23) was confirmed
by the above mentioned chemical analyses. The comparison
between the measured and calculated values of calcium sulfate
is presented in Fig. 6a. The experimental data were found to be
in good agreement with the calculated values. The mass balance
for ferruginous active absorbent was determined as shown in
Fig. 6b. Both figures demonstrated a fair mass balance.
2.5. Thermodynamics
Whether a chemical reaction is spontaneous or not is associated
with the standard-state Gibbs free energy,
Δr GΘm ¼
X
B
γB Δ f GΘm ðB; βÞ
ðE3Þ
where, ΔγGΘ
m (kJ/mol) is the standard-state Gibbs free energy;
Δf GΘ
m (kJ/mol) is the free energy of the reactant or product; and γB
is the stoichiometric ratio in the reaction. According to thermodynamics, the standard-state Gibbs free energies of the main
overall reactions (Reactions (21) and (23)) were determined to be
−575.57 and − 246.66 kJ/mol, respectively, which means that
the desulfurization and denitration using the ferruginous
active absorbent are exergonic and spontaneous.
The equilibrium constants can be further obtained by
ΔGΘ
K ¼ e− RT
ðE4Þ
where, K is the equilibrium constant; T (K) is the reaction
temperature; and R (R = 8.314 J/(K·mol)) is the universal gas
constant. Under standard conditions (298 K), the equilibrium
constants of Reactions (21) and (23) are calculated as 7.8E+100 and
1.73E+43, respectively. Such a large value of the equilibrium
constants demonstrates that the desulfurization and denitration
using the ferruginous active absorbent are positive and thorough.
At a different temperature, the partial pressures are calculated
as
ΔGΘ þ RT ln J p ¼ 0
ðE5Þ
Calculated calcium sulfate (g)
25
Jp ¼
Experimental data
Calculated values
20
10
5
a
0
5
10
15
20
25
Measured mass of Calcium sulfate (g)
Output exhaut gas plus accumulation and
comsumption of absorbent (g/m3)
ðE6Þ
where, pc and pa are the partial pressures of products and
reactants, respectively; Jp is the reaction quotient. The partial
pressures of the solid products in the system can be disregarded
(ln( pc) = 0). Substituting the values of standard-state Gibbs free
energy (−575.57 kJ/mol for desulfurization and −246.66 kJ/mol
for denitration), universal gas constant (8.314 J/(K·mol)) into
Eqs. (E5) and (E6), the partial pressures of SO2–NOx binary gases
can be obtained as shown in Eqs. (E7) and (E8), respectively,
without considering the impact of oxygen.
15
0
pc
Pa
23:1 103
T
ðE7Þ
4:94 103
T
ðE8Þ
ln pSO2 ¼ −
50
Input
Output+accumulation+consumption
ln pNO ¼ −
40
Thus, the partial pressures of SO2–NOx binary gases are
summarized in Table 3 at the five adopted temperatures. It
can be seen that the equilibrium partial pressures of SO2 and NO
rise with the reaction temperature, for which the increase is
particularly marked in desulfurization, with slight increases in
denitration. Such theoretical results are consistent with the
above experiments.
30
20
10
b
0
0
10
20
30
40
3. Conclusions
50
Input SO2-NO binary gas and fresh absorbent (g/m3)
Fig. 6 – (a) Comparison curves between measured and
calculated calcium sulfate; (b) mass balance of input total
amount of SO2–NO binary gas and loading fresh absorbent,
and output exhaust gas plus accumulation and consumption
of absorbent.
A novel ferruginous active absorbent was fabricated in this work.
The addition of Fe(VI) supplies oxidizing sites in the absorbent,
which can convert NO and SO2 into higher valence nitrogen and
sulfur compounds synchronously. The porous fly ash might play
an adsorptive and catalytic role during the Fe(VI) oxidative
removal of SO2 and NO. After the removal reaction, most of the
63
J O U RN A L OF E N V I RO N ME N TA L SC IE N CE S 3 0 (2 0 1 5) 5 5–6 4
Table 3 – Partial pressures of SO2-NOx binary gases.
Temperature (K)
pSO2 (MPa)
pNO (MPa)
323
328
333
338
343
9.383E −32
2.247E −7
2.788E−31
2.837E−7
8.019E−31
3.558E−7
2.235E−30
4.432E−7
6.047E−30
5.486E−7
solid products covered the surface of the absorbent in the form of
a “white flake layer”.
Experiments of synchronous desulfurization and denitration
were carried out using the prepared ferruginous active absorbent in a fixed bed reactor. The study of synergistic effects of
operating factors found that strong synergy exists between
Fe(VI) dose and reaction temperature for the desulfurization;
while in the denitration process, the synergy of Fe(VI) dose and
Ca/(S + N) has the most significant impact on the removal of
NO, followed by the synergy of Fe(VI) and reaction temperature,
and then the synergy of reaction temperature and flue gas
humidity.
The appearance of nitrogen and the increased level of sulfur
content in the energy spectrum of the spent absorbent indicates
that the removal reactions are mainly chemical absorption
between SO2, NO and the ferruginous active absorbent, in
other words, NO and SO2 were oxidized first and then absorbed
chemically. The stepwise reaction paths were analyzed. Calcium
sulfate and calcium nitrate were the main removal products of
desulfurization and denitration, respectively.
Acknowledgments
The research was supported by the National Natural Science
Foundation of China (Nos. 51308212, 10974053), the National Key
Technology R&D Program (No. 2011BAI02B03), the Fundamental
Research Funds for the Central Universities (No. 13ZD18), and the
State Scholarship Fund (No. 201206735009). The authors sincerely
thank the School of Energy Resources at the University of
Wyoming and the School of Civil and Environmental Engineering
at Georgia Institute of Technology for their great support and
technological guidance.
Appendix A. Supplementary data
Supplementary data to this article can be found online at
http://dx.doi.org/10.1016/j.jes.2014.10.012.
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