Environ. Sci. Technol. 2002, 36, 247-253
Biofiltration of Methyl tert-Butyl
Ether Vapors by Cometabolism with
Pentane: Modeling and
Experimental Approach
D A V I D D U P A S Q U I E R , †,‡
S E R G I O R E V A H , * ,† A N D
R I C H A R D A U R I A †,‡
Departmento de Ingenierı́a Quı́mica, Universidad Autónoma
Metropolitana, UAM-Iztapalapa, Apdo Postal 55-534,
CP 09340, México City, México, and Institut de Recherche
pour le Développement (IRD), Calle Cicerón 609,
Los Morales 11530, México City, México
Degradation of methyl tert-butyl ether (MTBE) vapors by
cometabolism with pentane using a culture of pentaneoxidizing bacteria (Pseudomonas aeruginosa) was studied
in a 2.4-L biofilter packed with vermiculite, an inert
mineral support. Experimental pentane elimination capacity
(EC) of approximately 12 g m-3 h-1 was obtained for an
empty bed residence time (EBRT) of 1.1 h and inlet
concentration of 18.6 g m-3. For these experimental
conditions, EC of MTBE between 0.3 and 1.8 g m-3 h-1
were measured with inlet MTBE concentration ranging from
1.1 to 12.3 g m-3. The process was modeled with general
mass balance equations that consider a kinetic model
describing cross-competitive inhibition between MTBE
(cosubstrate) and pentane (substrate). The experimental
data of pentane and MTBE removal efficiencies were
compared to the theoretical predictions of the model. The
predicted pentane and MTBE concentration profiles
agreed with the experimental data for steady-state operation.
Inhibition by MTBE of the pentane EC was demonstrated.
Increasing the inlet pentane concentration improved
the EC of MTBE but did not significantly change the EC of
pentane. MTBE degradation rates obtained in this study
were much lower than those using consortia or pure strains
that can mineralize MTBE. Nevertheless, the system can
be improved by increasing the active biomass.
Introduction
In many countries (United States, Mexico, France, etc.), fuel
oxygenates are added to gasoline in order to enhance its
octane number. In the United States, they have been used
since 1988 to improve air quality in some metropolitan areas
(1). These additives allow a better gasoline combustion and,
consequently, reduce the resulting concentrations of carbon
monoxide and unburned hydrocarbons. Among the oxygenates (ethanol, ethyl tert-butyl ether, tert-amyl methyl ether,
etc.), methyl tert-butyl ether (MTBE) is being used the most
by the industry because of its low cost, ease of production
at refineries, favorable blending characteristics with other
fuel components, and lack of phase separation in the presence
* Corresponding author phone: +52-55-58-04-65-38; fax: +5255-58-04-64-07; e-mail: [email protected].
† Universidad Autónoma Metropolitana.
‡ Institut de Recherche pour le Développement.
10.1021/es010942j CCC: $22.00
Published on Web 12/13/2001
 2002 American Chemical Society
of water. Between 1984 and 1994, MTBE production rates
increased by 26% annually in the United States. In 1993, its
production ranked second among all organic chemicals
manufactured (2). Consequently, the use of MTBE as an
additive in gasoline has been increasing and has extended
to other countries (Mexico, China, Saudi Arabia, etc.).
However, due to this massive production combined with its
mobility, persistence, toxicity, and high solubility in water
(50 g L-1 at ambient temperature), it is an important pollutant
of groundwater (3, 4). It has generally been considered that
the persistence of MTBE is strongly related to its biological
recalcitrance, although recent studies have shown that there
is strong potential for natural attenuation (5). Because of the
presence of MTBE in the environment, the U.S. EPA has
announced its intention to reduce or eliminate this oxygenate
in domestic fuels over the next 3 yr.
To date, very few studies have been conducted to find
microorganisms able to degrade MTBE and to utilize them
in biological treatment. Consortia and pure microorganisms
have been studied for their ability to degrade MTBE as the
sole source of carbon and energy in aerobic (6, 7) and
anaerobic conditions (8). However, a limited amount of work
has been performed on the cometabolic degradation of MTBE
by pure cultures. In one recent study, degradation of MTBE
by filamentous fungus in the presence of diethyl ether (DEE)
was reported (9). Biodegradation of MTBE by three propaneoxidizing strains (ENV421, ENV425, and Pseudomonas putida)
was studied (10). Garnier et al. (11) showed that a soil
consortium was able to degrade to completion gasoline
containing MTBE. However, when MTBE was tested alone,
no degradation was observed. This study demonstrated that
MTBE was degraded by cometabolism with n-alkanes (pentane, hexane, and heptane) present in the gasoline. From
the consortium, a pentane-oxidizing bacteria (P. aeruginosa)
isolated from this consortium was able to degrade MTBE by
cometabolism with pentane (12). Biofiltration of MTBE vapors
released from a wastewater treatment was reported (13).
Acclimation of the microorganisms for more than 1 yr was
necessary to observe MTBE degradation. In this case, the
biofilter performance was approximately 7 g m-3 h-1, which
corresponded to a removal efficiency of nearly 97%. After 6
months of acclimation of aerobic microbial consortium able
to degrade MTBE, Fortin and Deshusses (14) obtained
elimination capacity of MTBE around of 50 g m-3 h-1 (90%
of removal efficiency) in laboratory-scale biotrickling filters.
A pure strain, with high MTBE mineralization rate, was
reported (15). This strain nevertheless shows some inhibition
when BTEX is present (16). For more information, the reader
will be able to refer to the paper by Deeb et al. (17) on aerobic
MTBE biodegradation.
Biological waste air treatment has proven to be a good
alternative as compared with conventional treatments (stripping, incineration, adsorption, absorption, ...). Until 1980,
biofiltration was mainly used to reduce odor in off-gases,
but in the early 1980s, the field of application was extended
to the removal of many other volatile organic compounds.
Biofiltration is a cost-effective method to treat large volumes
of contaminated air with moderate concentrations of volatile
organic compounds (18). This technology has low power
requirements, and the process equipment is simple and
generally easy to operate. In biofiltration, the gas to be treated
is forced through a bed packed with material on which
microorganisms are attached as a biofilm. Biodegradable
volatile compounds and oxygen diffuse into the biofilm where
they are subsequently biologically oxidized into less harmful
substances such as CO2 and H2O.
VOL. 36, NO. 2, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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The aim of this study is to investigate the degradation of
MTBE vapors by cometabolism with pentane using a culture
of pentane-oxidizing bacteria. The study was carried out with
a biofilter packed with a mineral support (vermiculite). A
mathematical model was proposed to understand and predict
cometabolic degradation of MTBE in the presence of pentane
in the reactor. Its predictions were compared with experimental results obtained from the biofilter, and a sensitivity
study of the different model parameters was conducted.
gbiomass-1 s-1, Xfa is the active biomass concentration in g
mbiolayer-3, and Ks(i) is the half-saturation constant of compound i in g m-3.
MTBE is not consumed as a source of carbon for cell
synthesis (11); however, its degradation is possible when
pentane is supplied in the reactor. Therefore, the biodegradation of MTBE occurs only in the presence of pentane.
MTBE degradation is described by eq 3, considering the
competitive inhibition and the stimulating effect of pentane:
Theory
Sm
rm ) kx(MTBE)Xfa
Sm + Ks(MTBE) 1 +
The theoretical model presented in this section was based
on that proposed by Ottengraf and van den Oever (19). Crosscompetitive inhibition between MTBE (cosubstrate) and
pentane (substrate) was described using a kinetic model
similar to the one presented by Arcangeli and Arvin (20).
To predict the degradation of MTBE and pentane in the
biofilter, the following assumptions were made:
(i) The airstream passes through the biofilter bed in plug
flow mode.
(ii) The gas pollutants at the biofilm/air interface are in
equilibrium as dictated by Henry’s law.
(iii) The thickness of the biofilm (δ) is small as compared
to the radius of the particles, and thus, planar geometry can
be used.
(iv) Pollutants are transported by diffusion in the biofilm.
(v) The biofilm is homogeneous, and its density, defined
as the amount of dry biomass per unit volume of biolayer,
is constant.
(vi) Only pentane (substrate) and MTBE (cosubstrate) are
assumed to be rate limiting.
The differential equation describing the concentration of
the compound i (Si) inside the biolayer at steady-state
conditions is
d2Si
Dei 2 ) ri
dx
(1)
where Dei is the diffusion coefficient of compound i into the
biofilm (m2 s-1); ri is the reaction rate of product i (g m-3 s-1);
and i is pentane (p) or MTBE (m). The following boundary
conditions are employed:
[
x ) 0 Si ) Cgi/mi
x ) λ dSi/dx ) 0
]
(
Sm
Ks(MTBE)
)
(2)
where kx(i) is the maximum substrate utilization rate in gi
248
9
(
Ks(pent)
Sp
)
×
Sp + Ks(pent)
)
(3)
The kinetic model considers that the total biomass (Xf) is
comprised of two parts: an active biomass (Xfa) capable of
degrading pentane and MTBE and an inactive biomass, which
includes dead cells and exopolymeric substances (EPS) (18,
19).
From a differential mass balance in the gas phase, Cgi
may be calculated as a function of the height (h) in the filter
bed according to
-Ug
dCgi
) NiAs
dh
(4)
where Ni (g m-2 s-1) is the substrate flux into the biofilm, As
is the interfacial area per unit of reactor volume (m2 m-3);
h is the distance from the biofilter entrance (m); and Ug is
the superficial gas flow rate (m s-1). Finally, the expression
for Ni is given by the Fick’s law:
( )
Ni ) - Dei
dSi
dx
x)0
(5)
A Runge-Kutta integration algorithm was used to solve the
system of eqs 1-5. The integration step in biofilter thickness
(∆δ) and height (∆h) was equal to 0.1. This model estimated
the different concentration profiles for each pollutant in the
biofilm and along the biofilter.
Materials and Methods
where Cgi is the gas concentration of compound i (g m-3), λ
is the active biofilm thickness (m), and mi is the distribution
coefficient for compound i in the air/water system.
The second boundary condition follows from the consideration that somewhere in the biofilm the concentration
gradient equals zero. In the reaction-limiting case, the
condition is met at x ) λ ) δ. When diffusion limitation
exists, the depth of penetration λ in the biofilm may be smaller
than its thickness δ; however, the condition is still valid.
The reaction rate of pentane, which supports biomass
growth, can be described by eq 2, assuming a competitive
inhibitory effect of the cosubstrate. The inhibition coefficient
of the competitive inhibitor is approximated by its singlesubstrate half-saturation coefficient. Although an inhibitory
effect of pentane has been found for this system (12), it was
not considered as the gas-phase concentrations used in this
study were much lower than those used in ref 12:
Sp
rp ) kx(pent)Xfa
Sp + Ks(pent) 1 +
(
Sp
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 36, NO. 2, 2002
Inoculum and Mineral Medium. A bacteria (P. aeruginosa),
isolated from soil samples contaminated with gasoline (12),
was used for the experiments. This bacteria was able to
degrade MTBE when cometabolized in the presence of
pentane. The isolated strain was enriched, and 2 L of medium
was prepared for the inoculation. The medium was a mineral
salt solution consisting of the following components: MgSO4‚
7H2O, 1.0 g L-1; KNO3, 1.0 g L-1; CaCl2, 0.2 g L-1; Fe-EDTA,
0.4 g L-1; K2HPO4, 74 mg L-1; KH2PO4, 26 mg L-1; FeSO4, 0.2
mg L-1; ZnSO4‚7H2O, 0.01 mg L-1; H3BO3, 0.03 mg L-1; CoCl2‚
2H2O, 0.02 mg L-1; MnCl2‚4H2O, 0.003 mg L-1; NiCl2‚2H2O,
0.002 mg L-1; CaCl2, 0.001 mg L-1; NaMoO4, 0.03 mg L-1.
Inoculation was made by flooding the entire reactor twice
and keeping each time the solution for 30 min in contact
with the support before pumping it out from the reactor.
Chemicals. MTBE (98%, d ) 0.740) was from Sigma (St.
Louis, MO). A pressurized cylinder containing gaseous 1.0%
pentane in air (Praxair, México) was used to feed the substrate.
Biofilter and Packing Material. A three-stage biofilter
consisted of a cylindrical glass column with i.d. of 0.078 m
and 1 m height. It was equipped with a number of sampling
ports along its length (Figure 1). Each filter stage was packed
with a mineral support (vermiculite) and supported by a
plate to ensure a homogeneous packing distribution. Particles
of vermiculite were sieved through -4+5 mesh screens to
TABLE 1. Operating Conditions of the Biofilter
inlet MTBE concn (g m-3)
inlet pentane concn (g m-3)
air flow (L h-1)
reactor vol (L)
temp (°C)
EBRT (τ) (h)
1.1-12.3
0.7-19.2
0.8-39.5
2.4
30
0.06-2.85
tration difference (Cgin,i - Cgout,i) and the inlet concentration
(Cgin,i).
Determination of Model Parameters
FIGURE 1. Biofilter system: 1, air compressor; 2, pentane tank; 3,
biofilter; 4, mass flow controller; 5, unidirectional valve; 6, humidifier;
7, cyclone; 8, mixing bottle; 9, MTBE bottle in a refrigerated system;
10, peristaltic pump.
obtain an average particle size of 4.1 mm. The active filter
bed height was 0.5 m, and the packing density was 110 g of
dry vermiculite L-1. The initial water content and pH were
70% and 7.0, respectively. The biofilter was placed in a
controlled chamber at 30 ( 2 °C.
Addition of MTBE and Pentane. A stream containing 1%
pentane in air from a pressurized cylinder was supplied
directly to the reactor after being saturated with water vapor
by sparging through a column. Pentane concentrations lower
than 1% were obtained by mixing pentane with compressed
air. The main airstream was controlled by an electronic mass
flow sensor (33116-20 Cole Parmer, USA). A small air flow
was sparged through a 0.5-L flask containing liquid MTBE
and mixed with the air contaminated by pentane before
entering at the top of the reactor. The MTBE flask was placed
into a refrigerated system to control temperature variations.
Analysis. Pentane and MTBE concentrations in the gas
phase were determined by gas chromatography. A 250-µL
airtight syringe was used to sample at different heights in the
biofilter. These samples were injected into a FID gas
chromatograph (Hewlett-Packard 5890, USA) equipped with
a silica capillary column (30 m CP WAX 52 CB, USA). The
operating conditions were as follows: injector, 225 °C; oven,
40 °C; and detector, 225 °C. The detection limit was
approximately 5 mg of pollutant/m3 of air.
The Lowry method was used to determine total biomass
concentration assuming that 50% of total biomass is protein.
Active biomass concentration was estimated from the work
of Pineda et al. (21).
Experimental Procedure. The first series of experiments
were conducted feeding only pentane to the biofilter. These
experiments were performed with different pentane inlet
concentrations (Cgin(pent)) ranging from 0.72 to 17.4 g m-3,
and empty bed residence times (EBRT ) τ) range from 0.06
to 2.85 h. After steady state was attained, air samples were
taken from different levels of the column. The second series
of experiments were conducted in the presence of pentane
and MTBE. These experiments were carried out by maintaining Cgin(pent) (18 g m-3) and τ (1.1 h) constant. Only, the
inlet MTBE concentration (Cgin(MTBE)) was varied from 1.1
to 12.3 g m-3. Operating conditions of the biofilter are
presented in Table 1.
Definitions. The elimination capacity (EC) of compound
i is defined as the difference between inlet (Cgin,i) and outlet
(Cgout,i) concentrations divided by the EBRT (τ). EC is
expressed as gpollutant m-3biofilter h-1. Removal efficiency (%Ef)
of compound i is defined as the ratio between the concen-
To determine As, δ, Xf, and Xfa, the model of biolayer growing
on the vermiculite support and the experimental methods
presented by Pineda et al. (21) were used. To determine As,
the form of the vermiculite particles is assumed to be cubic.
These cubes are formed by layered and sandwiched thin
mineral sheets. Observations by scanning electronic microscope (SEM) showed that the formation of the biofilm takes
place between these layers and that the biofilm only grew on
four faces of the particles. The volume occupied by the
vermiculite was Vv ) (1 - )Vt, where is the porosity of the
packing material ( ) 0.71) and Vt is the volume of the reactor
(2.4 L). Since the vermiculite particles are cubic, the volume
of one particle was V1 ) a3 (where a ) 4.1 mm). Therefore,
the approximate number of particles (n) constituting the
packing material is equal to Vv/V1. Considering that the
exchange area for one particle is A1 ) 4a2, the total exchange
area was At ) nA1. Finally, As ) At/Vt was equivalent to As )
284 m2 m-3 of reactor for this study.
Pineda et al. (21) estimated the volume of superficial water
by differential thermogravimetry. They considered that this
superficial water is the first to evaporate and therefore
corresponds to the first peak of the differential thermogravimetry spectrum of wet vermiculite. The volume of
superficial water was considered equal to the biofilm volume.
The superficial water represented about 25% of the total
water, which is equal to 1 g of water/g of dry vermiculite.
Given the dry packing material density, the total volume of
water (Vw) constituting the biolayer was determined (Vw )
259.6 cm3). Since the thickness of the biolayer is defined by
δ ) Vw/(AsVt), the numerical value was found to be δ ) 387
µm.
The total biomass density of the biofilm (Xf) was equal to
28 300 g m-3. The active biomass was considered equal to
8% of total biomass (21). Thus, the active biomass density
of the biofilm (Xfa) was calculated to be 2264 g m-3. In the
literature, a wide range of values for As (25-1000 m-1) is
found for different supports as reviewed by Pineda et al. (21).
In the case of vermiculite particles, an As value of 675 m-1
was reported for a size distribution ranging from 0.52 to 5.4
mm (21). The value of As reported here for the same packing
material is lower since a larger mean particle size (4.1 mm)
was used. Values of δ (387 µm) and Xf (28 300 g m-3),
calculated in this study, are in the same order of magnitude
as those reported by different authors (21). From these values,
a total active biomass in the reactor can be calculated as
δAsXfa ) 250 g mreactor-3, which was lower than values that
have been reported elsewhere (2, 18).
Diffusivities of pentane and MTBE (Dei) in the biolayer
were assumed to be equal to their diffusivities in water (Di),
corrected by a factor depending on the Xf, according to the
expression of Fan et al. (22).
Kinetic parameters, Ks(pent), Ks(MTBE), kx(pent), and kx(MTBE) were previously determined from batch experiments
(12). The half-saturation constant of pentane (Ks(pent)) was
equal to the apparent saturation constant deduced from the
integrated form of the first-order growth equation coupled
with Monod kinetics. This value (0.019 g m-3) was obtained
VOL. 36, NO. 2, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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249
TABLE 2. Values of the Model Parametersa
De(pent) (m2 s-1)
De(MTBE) (m2 s-1)
kx(pent) (g g-1db s-1)
kx(MTBE) (g g-1db s-1)
Ks(pent) (g m-3)
Ks(MTBE) (g m-3)
mpent at 25 °C
mMTBE at 25 °C
h (m)
δ (m)
As (m2 m-3)
Xf (g db m-3biofilm)
Xfa (g db m-3biofilm)
T (°C)
a
1 × 10-9
0.9 × 10-9
58.4 × 10-6
2.87 × 10-6
0.019
185
44.4
0.022
0.5
0.000387
284
28,300
2,264
30
db, dry biomass.
for a pentane concentration of 2.9 µg L-1 and represents the
real saturation constant of pentane. The pentane maximum
substrate utilization rate (kx(pent)) was obtained by dividing
the maximum growth rate (µmax ) 0.19 h-1) by the growth
yield coefficient (Yg ) 0.9 g of biomass/g of pentane). MTBE
maximum substrate utilization rate (kx(MTBE)) was calculated from the value of the MTBE degradation rate (3.9 nmol
min-1 mg-1cell protein) using the MTBE molecular weight and
percentage of protein of dry biomass, 88.15 g mol-1 and 50%,
respectively. Table 2 presents the parameter values used for
solving the model equations.
Few studies have addressed the cometabolic biodegradability of MTBE by consortia and pure cultures. Hardison et
al. (9) reported that MTBE was degraded by a filamentous
fungus (Graphium sp.) in the presence of butane at a ratio
varying between 0.03 × 10-6 and 0.252 × 10-6 g g-1db s-1.
Biodegradation of MTBE by three propane-oxidizing strains
with rates between 0.3 × 10-6 and 6.7 × 10-6 g g-1db s-1 were
reported (10). These values are in the same order of magnitude
as found in this study (kx(MTBE) ) 2.87 × 10-6 g g-1db s-1).
However, Hanson et al. (15) obtained higher kx(MTBE) values
(up to 1.8 × 10-3 g g-1db s-1) with a bacterial pure culture that
utilized MTBE as its sole carbon and energy source. No data
of Ks(pent), Ks(MTBE), and kx(pent) pertaining to the cometabolism of MTBE with alkanes are available in the literature.
Nevertheless, values of similar parameters for TCE (trichloroethylene) degraders by cometabolism growing on toluene
as the sole carbon sources were reported (20). The values of
Ks(toluene) and kx(toluene) varied between 0.027 and 13.8
g m-3 and between 2.4 × 10-6 and 1.9 × 10-4 g g-1db s-1,
respectively. These values are in agreement with these
obtained in this study for Ks(pent) and kx(pent) (Table 2).
Arcangeli and Arvin (20) reported values of Ks(TCE) (0.173
g m-3 < Ks(TCE) < 19.6 g m-3) lower than Ks(MTBE) (185 g
m-3). This difference probably is due to the lower affinity for
MTBE as compared to TCE for the respective enzyme
produced when pentane and toluene are used as carbon
sources.
Results and Discussion
Degradation of Pentane without MTBE Addition. Figure
2a-c represents the model and experimental profiles of
pentane vapor concentrations along the column obtained
for different conditions of EBRT and inlet pentane concentrations (Cgin(pent)). These experiments show that the model
gives a good representation of pentane degradation. Better
agreement was found for Figure 2b,c where the experimental
errors were reduced. For Figure 2a,b, the model predicts a
linear decrease in pentane concentration in the filter bed,
which suggests that the biofilm was fully active. In this case,
the degradation rate is independent of the substrate concentration and the degradation in the biofilm follows zero250
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 36, NO. 2, 2002
FIGURE 2. Predicted and experimental profiles of the pentane
concentration in the biofilter for pentane inlet concentration of 0.72
g m-3 and τ of 0.077 h (a); 17.4 g m-3 and τ of 0.58 h (b); 13.5 g m-3
and τ of 2.85 h (c). Experimental values (b) and model predictions
(s). Error bars (|).
TABLE 3. Experimentally Obtained and Model-Predicted
Steady-State Elimination Capacity for Pentane Vapors Alonea
τ in h
Cgin(pent)
(g m-3)
ECexp
(g m-3 h-1)
ECpred
(g m-3 h-1)
error
(%)
0.0583
0.077
0.58
2.11
2.85
2.4
0.7
17.4
14.9
13.5
6.8
1.3
12.0
7.0
4.7
2.7
0.8
12.2
6.5
4.6
-60.3
-36.4
+1.3
-7.3
-1.0
a EC
exp, experimental elimination capacity; ECpred, predicted elimination capacity.
order kinetics (23). At high EBRT (τ ) 2.85 h) (Figure 2c), the
biofilter was not completely active. The model predicts a
non-linear decrease in pentane concentration in the biofilter
that suggests that the biofilter zones are limited by diffusion.
Table 3 presents the percentage of error obtained from
the comparison between the mathematical model and the
experimental data for different Cgin(pent) and EBRT. The
highest percentage of error, ranging from - 60 to - 36%,
were obtained for the low Cgin(pent) and EBRT while the errors
(-7.3 to + 1.3%) decreased when Cgin(pent) and EBRT are
higher. In these experiments, the highest pentane %Ef (40%)
was obtained at the maximum EC of 12 g m-3 h-1.
n-Alkanes are strongly hydrophobic, having Henry’s
coefficients (m) in the range of 10-100 (g m-3gas/g m-3water).
Because of poor mass transfer from the gas to the aqueous
phase, low ECs for n-alkanes were measured in biofilters.
For example, low hexane ECs of (0.2-5.4 g of methane equiv.
m-3 h-1) were attained using biofilter treating fuel vapors
(24). In a full-scale biofilter treating hexane from waste gases
of an oil refinery, 68% was eliminated at an EC of 2.5 g m-3
h-1 (25). Higher ECs of hexane were reported in laboratoryscale biofilters: 32 g m-3 h-1 (%Ef ) 39%) (25) and 21 g m-3
h-1 (%Ef ) 99%) (26).
TABLE 4. Elimination Capacities of Pentane and MTBE Obtained from Mathematical Model and Experimentsa
τ in h
Cgin,p
(g m-3)
Cgin,m
(g m-3)
ECexp,p
(g m-3 h-1)
ECpred,p
(g m-3 h-1)
error,p
(%)
ECexp,m
(g m-3 h-1)
ECpred,m
(g m-3 h-1)
error,m
(%)
1.1
1.1
1.1
1.1
1.1
17.9
19.2
18.8
18.8
18.7
1.1
4.8
9.6
11.0
12.3
9.1
7.2
7.5
5.6
4.4
9.7
7.5
5.3
5.1
4.5
+7.6
+3.7
-29.5
-8.5
+ 2.3
0.3
0.7
1.5
0.7
1.8
0.2
0.6
0.8
0.9
0.9
-45.4
-13.2
-45.4
+23.7
-47.8
a
ECexp, experimental elimination capacity; ECpred, predicted elimination capacity.
FIGURE 4. Influence of the inlet MTBE concentration (Cgin) on the
elimination capacities (EC) of pentane and MTBE. Experimental
values: (b) pentane, (O) MTBE. Model predictions: (s) pentane,
(- -) MTBE.
FIGURE 3. Predicted and experimental profiles of the pentane and
MTBE concentrations in the biofilter. (a) Pentane inlet concentration
of 19 g m-3 and MTBE inlet concentration of 4.8 g m-3. (b) Pentane
inlet concentration of 18 g m-3 and MTBE inlet concentration of 11
g m-3. Experimental values: (b) pentane, (O) MTBE. Model
predictions: (s) pentane, (- -) MTBE. Error bars (|).
This first series of experiments showed that the model
predicted well the experimental profiles of pentane concentration along the biofilter and, consequently, that the
hypotheses used for the determination of the As, δ, Xf, Xfa,
and Ks(pent) were appropriate. It should be noted that a
preliminary fitting of these parameters with the mathematical
model was not used.
Degradation of MTBE by Cometabolism with Pentane.
A second series of experiments were conducted by maintaining the inlet pentane concentration (18 g m-3) and the
EBRT (1.1 h) constant. Only the inlet MTBE concentration
was varied as shown in Table 1.
Figure 3a,b represents pentane and MTBE concentration
profiles along the biofilter for inlet MTBE concentrations
(Cgin(MTBE)) of 4.8 and 11 g m-3. At the steady-state condition,
a low MTBE EC of 0.7 g m-3 h-1 was found for both
experiments. Experiments in the biofilter (not shown) showed
that the MTBE was not degraded in the absence of pentane.
The EC of MTBE was much lower than the ECs reported in
biofilter (ECmax ) 15 g m-3 h-1) (13) and trickling biofilter
(ECmax ) 50 g m-3 h-1) (14). Both works have reported that
MTBE is completely mineralized by their consortia. For
pentane, the EC was maintained at approximately 7 g m-3
h-1. For the two compounds, the relation between the
concentration and the biofilter height was linear. This
behavior indicates that rates of MTBE and pentane degradation were limited by biological reaction. This can be further
supported by the fact that the system operated with low
biomass as mentioned earlier. Improved volumetric rates
can be obtained by increasing the biomass in the reactor
through nutrient medium addition or by the use or more
efficient strains. The mathematical model predicts, with
reasonable error (Table 4), the experimental profiles of the
pentane and MTBE concentration. Errors in the pentane and
MTBE EC prediction ranged from 3.7% to -13.2% (Figure
3a) and from -8.5% to 23.7% (Figure 3b).
The studies of Garnier et al. (12) showed in microcosms
the effect of the concentration of both compounds on the
elimination rates. The effect of the Cgin(MTBE) on the EC of
pentane and MTBE was investigated (Figure 4). Higher EC
of MTBE were found as the inlet MTBE concentration
increases. On the other hand, the marked inhibition of MTBE
on the EC of pentane appears very clearly in this figure (50%
at around 10 g m-3). Moreover, as the ratio of pentane/MTBE
decreases, the EC of MTBE increases slowly. The inhibition
of MTBE on pentane degradation is confirmed from the
results of the Figure 5, which represents %Ef of pentane and
MTBE as a function of the MTBE inlet concentration. When
the concentration of MTBE increases from 0 to 20 g m-3, %Ef
of pentane and MTBE decrease from 60 to 20% and from 20
to 6%, respectively. Similarly, the %Ef of MTBE decreases but
more slowly. The competitive inhibition of MTBE on the
pentane uptake was detected even at the lowest MTBE
concentrations. A small inhibitory effect of liquid MTBE
concentrations of 200 mg L-1 on the biodegradation of
aromatic hydrocarbons (benzene, toluene, ethylbenzene, and
xylenes) was reported (27). Arcangeli and Arvin (20) investigated the cometabolic degradation of TCE with toluene as
the primary source of carbon in a continuously fed biofilm
reactor. These authors showed that TCE inhibits toluene
degradation for TCE concentrations above 50 µg/L. These
results were confirmed from an investigation of TCE degVOL. 36, NO. 2, 2002 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 5. Model Sensitivity to Values of Pentane Half-Saturation Constant (Ks(pent)), MTBE Half-Saturation Constant (Ks(MTBE)),
Pentane Maximum Substrate Elimination (kx(pent)), MTBE Maximum Substrate Elimination (kx(MTBE)), Interfacial Area Per Unit of
Reactor Volume (As), Total Biomass Concentration (Xf), and Biofilm Thickness (δ)a
E (%)
error on
R pentane (%)
error on
R MTBE (%)
Ks(pent)
Ks(MTBE)
kx(pent)
kx(MTBE)
As
Xf
δ
(20
(+) -13.5
(-) +16.7
(+) -9.2
(-) +10.6
(20
(+) +6.3
(-) -9.0
(+) -10.0
(-) +12.2
(20
(+) +15.1
(-) -16.5
(+) -2.8
(-) +2.8
(20
(+) +0.5
(-) -0.5
(+) +20
(-) -20
(20
(+) +15.0
(-) -18.0
(+) +15.0
(-) -18.0
(20
(+) +15.0
(-) -18.0
(+) +15.0
(-) -18.0
(20
(+) +15.0
(-) -18.0
(+) +15.0
(-) -18.0
a R is the relative value of elimination capacity. The reference values of pentane and MTBE EC are 5.95 and 0.52 g m-3 h-1, respectively. These
values were obtained with a Cgin(pent) ) 15 g m-3, Cgin(MTBE) ) 5 g m-3, and τ ) 1 h. The reference value for Ks(pent), Ks(MTBE), kx(pent), kx(MTBE),
As, Xf, and δ are those reported in Table 2. (+)/(-), represents the sign of the percentage variation of Ks(pent), Ks(MTBE), kx(pent), kx(MTBE), As,
Xf, and δ.
FIGURE 5. Influence of the inlet MTBE concentration (Cgin) on the
removal efficiencies (%Ef) of pentane and MTBE. Experimental
values: (b) pentane, (O) MTBE. Model predictions: (s) pentane,
(- -) MTBE.
radation with P. putida F1 in the presence of toluene as the
primary carbon source (28). In contrast with MTBE, the
degradation of TCE may bring about the accumulation of
toxic intermediates. For the case of MTBE and pentane,
Garnier et al. (12) found for this bacterium that the ratio
between the real saturation constants, considering the
Henry’s coefficient, of MTBE and pentane was approximately
64 000. The much higher affinity for the pentane by the
enzymes involved in the degradation process was confirmed
in the biofiltration experiments.
Table 4 presents a comparison between the predicted EC
(ECpred) and experimental EC (ECexp) of pentane and MTBE.
Except for the Cgin(MTBE) of 9.6 g m-3, the percentage of
error of pentane and MTBE EC predicted ranged from -8.5
to +7.6%. These errors are higher (-45.4 to + 23.7%) in the
case of MTBE due to its low EC, which increases the
experimental error.
Figure 6 represents the influence of the Cgin(pent) on the
relative EC of pentane and MTBE obtained from the
mathematical model. These results were obtained by maintaining Cgin(MTBE) and the EBRT equal to 5 g m-3 and 1.1
h, respectively, and are reported as a percentage of the EC
(ECr) that would be obtained without MTBE (i.e., no
competitive inhibition). It is predicted that the ECr of pentane
varies little with the Cgin(pent) and decreases slowly when its
concentration increases. Up to a pentane concentration in
air of 20 g m-3 (0.45 mg L-1), no inhibition on MTBE
degradation was predicted (Figure 6) due to the low solubility
of pentane in water. This result is similar to that obtained
from batch experiments (11). These authors demonstrated
that the degradation of MTBE was stimulated in the presence
of pentane and proportional to its concentration. In the case
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FIGURE 6. Model predictions of the influence of the inlet pentane
concentration (Cgin(pent)) on the relative elimination capacities (ECr)
of pentane (s) and MTBE (- -) for Cgin, MTBE of 5 g m-3, and τ
) 1.1 h. ECr is reported as percent to the EC pent or MTBE obtained
without MTBE (i.e., no competitive inhibition).
of TCE degradation in cometabolism with toluene, Arcangeli
and Arvin (20) observed that degradation of TCE was inhibited
when toluene was present in concentrations above 1 mg L-1.
By differentiation of eq 3 with respect to Sp, an optimum
primary substrate concentration Sp opt, at which the cosubstrate (MTBE) degradation is maximized, can be calculated
(20). For a Cgin(MTBE) of 5 g m-3, which corresponds to a
concentration of MTBE in the biofilm of 227.3 mg L-1 (mMTBE
) 0.022 at 25 °C), a Sp opt of 1.27 mg.L-1 of pentane in the
biofilm was found. Then, in this case, a maximum efficiency
of MTBE degradation should be attained when the concentration of pentane in the gas phase is approximately to 56
g m-3 (for mpent ) 44.4 at 25 °C). For this concentration of
pentane, a low inhibition effect on the EC of pentane of 15%
is expected.
Sensitivity of Model. A sensitivity test was performed on
this model by modifying each parameter independently with
a (20% variation. Table 5 shows the results of the sensitivity
studies to the kinetics (Ks(pent), Ks(MTBE), kx(pent), and kx(MTBE)) and physical (As, Xf, and δ) parameters. All of these
sensitivity studies were conducted following the same
procedure.
In the Table 5, the error (E) corresponds to the relative
values of the kinetics and physical parameter divided by its
corresponding value (Table 2). Relative values (R) of the EC
of pentane and MTBE are defined as the EC corresponding
to the new values of the parameters divided by the EC
obtained from the mathematical model when Cgin(pent) )
15 g m-3, Cgin(MTBE) ) 5 g m-3, and EBRT ) 1 h. For these
conditions, the values obtained from the model for EC of
pentane and EC of MTBE were 5.95 and 0.52 g m-3 h-1,
respectively. It should be noted that since the comparisons
were based on the same EBRT and on the same inlet
concentrations, the relative EC (R) can be also viewed as the
relative percent of %Ef.
For a variation of (20%, the values of the affinity constants
affect both compounds similarly. On the contrary, variations
in the rate values (kx(pent) and kx(MTBE)) have a stronger
impact on their respective compounds. The result of the
sensitivity studies with three physical parameters (As, Xf, and
δ) have the same effect on the removal rate of each
compound. This result could be explained by the fact that
these three parameters were not characteristics of either of
the contaminants but were characteristics of the biofilter in
general. The relationship appeared quasi linear, and so an
increase of one parameter produces the increase of the
removal rate, and vice versa. This result would appear to be
normal and is confirmed by the form of the expression of EC
established by Ottengraf et al. (23) for a simpler metabolism
in the case of zero-order kinetics with reaction rate limitation
(EC ) kxAsδXf). However, in this study, As, δ, and Xf parameters
were not determined independently, and a variation of one
will affect the others. As shown in the sensitivity analyses,
to have appropriate kinetic and physical parameters for
biofiltration purposes, relatively accurate determinations are
needed to minimize errors.
Acknowledgments
D.D. was a visiting student to the UAM from the ENSCP
(Ecole Nationale Supérieure de Chimie Paris), France. The
authors appreciate the assistance of Armando Dreyer, David
Campos Santillan, Patrice Garnier, and Sergio Hernández.
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Received for review May 4, 2001. Revised manuscript received October 4, 2001. Accepted October 15, 2001.
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