MTBE oxidation by conventional ozonation and the combination ozone/hydrogen peroxide: Efficiency of the processes and bromate formation

Article abstract of DOI:10.1021/es010044n

The conventional ozonation and the advanced oxidation process (AOP) O₃/H₂O₂ was used to study the behavior of MTBE and its primary degradation products during ozonation and the AOP O₃/H₂O₂ for various raw waters. The major degradation products identified were tert-butyl formate (TBF), tert-butyl alcohol, 2-methoxy-2-methyl propionaldehyde (MMP), acetone, methyl acetate (MA), hydroxyisobutyraldehyde (HiBA), and formaldehyde. The rate constants of the reaction of O₃ and OH radicals with MTBE were 0.14 and 1.9 × 10⁹/M-sec, respectively. The rate constants for the same oxidation processes were also measured for the degradation products, TBF, MMP, MA, and HiBA. Since all compounds reacted slowly with molecular ozone, only the degradation pathway of MTBE with OH radicals was determined, including the formation of primary degradation products. In experiments conducted with several natural waters, the efficiency of MTBE elimination and bromate formation as disinfection byproduct were measured. With a bromide level of 50 μg/L, only 35-50% of MTBE could be eliminated by the AOP O₃/H₂O₂ without exceeding the current drinking water standard of bromate (10 μg/L). The transient concentrations of MTBE and its primary degradation products were modeled using a combination of kinetic parameters (degradation product distribution and rate constants) together with the ozone and OH radical concentration and were in good agreement with the experimental results.

Full text of DOI:10.1021/es010044n

Environ. Sci. Technol. 2001, 35, 4252-4259
tions), has a high water solubility (50 g/ L), and is mobile in
MTBE Oxidation by Conventional
Ozonation and the Combination
Ozone/Hydrogen Peroxide: Efficiency
of the Processes and Bromate
Formation
the subsurface (2). There is an ongoing scientific and public
debate on safe drinking water standards for MTBE. Currently,
there is no regulated value in place. However, the U.S. EPA
has issued a consumer acceptability advisory on MTBE with
a maximum tolerable concentration range in drinking water
of 20-40 µg/ L (on the basis of taste and odor considerations).
Recently, the California Department of Health Services
adopted a secondary maximum contaminant level of 5 µg/ L
for MTBE (3).
J U A N L . A C E R O , † S T E F A N B . H A D E R L E I N ,
T O R S T E N C . S C H M I D T ,
M A R C J . - F . S U T E R , A N D
Appropriate treatment options for MTBE-contaminated
raw water at waterworks are therefore necessary. The use of
air stripping, advanced oxidation processes (AOPs), and
adsorption to granular activated carbon or synthetic resins
has recently been compared in a comprehensive study with
emphasis on cost-effectiveness (4). The performance of all
methods depends on site-specific conditions such as water
flow, observed influent, and required effluent concentration.
However, MTBE is more difficult to remove from contami-
nated waters than other gasoline components when using
the more traditional treatment techniques (4). AOPs, which
are characterized by the production of hydroxyl radicals
(•OH), are considered a promising not yet fully established
treatment option for MTBE-contaminated water (4). Several
studies have been performed considering ozone/ hydrogen
peroxide (5, 6), UV/ hydrogen peroxide (7-9), Fenton’s
reagent (10), UV/ TiO2 slurries (11), and sonolytic destruction
in the presence of ozone (12, 13). The objectives of these
studies were to investigate the efficiency of MTBE degradation
by different AOPs and to identify the degradation products
formed. The major degradation products of MTBE are tert-
butyl formate (TBF), tert-butyl alcohol (TBA), acetone (AC),
and formaldehyde (FA) (5, 8, 10-12). In addition, 2-methoxy-
2-methyl propionaldehyde (MMP), hydroxyisobutyraldehyde
(HiBA), and methyl acetate (MA) were identified as primary
degradation products in some investigations (8).
Although a general reaction mechanism of MTBE deg-
radation by OH radicals has been elucidated, there is still a
lack of information about the relative contribution of the
different degradation pathways and, therefore, the distribu-
tion of intermediates during a given AOP treatment. More-
over, although some investigations have included the deg-
radation of MTBE in natural waters (6, 9), the influence of
drinking water parameters such as pH, alkalinity, natural
organic matter (NOM), and concentration of other interfering
compounds (e.g., bromide) has not been investigated. In the
present study, we focus on ozone-based oxidation processes
for MTBE degradation in natural waters. The impact of water
quality parameters on MTBE oxidation efficiency and the
formation of degradation products and bromate were evalu-
ated.
U R S V O N G U N T E N *
Swiss Federal Institute for Environmental Science and
Technology (EAWAG), Ueberlandstrasse 133,
CH-8600 Du¨ bendorf, Switzerland
The present study investigates the oxidation of methyl
tert-butyl ether (MTBE) by conventional ozonation and the
advanced oxidation process (AOP) ozone/hydrogen
peroxide under drinking water treatment conditions. The
major degradation products identified were tert-butyl formate
(TBF), tert-butyl alcohol (TBA), 2-methoxy-2-methyl
propionaldehyde (MMP), acetone (AC), methyl acetate
(MA), hydroxyisobutyraldehyde (HiBA), and formaldehyde
(FA). The rate constants of the reaction of ozone and OH
radicals with MTBE were found to be 0.14 and 1.9 × 109 M-1
s-1, respectively. The rate constants for the same
oxidation processes were also measured for the degradation
products TBF, MMP, MA, and HiBA (kO -TBF ) 0.78 M-1
3
s-1; kOH-TBF ) 7.0 × 108 M-1 s-1; kO MMP ) 5 M-1 s-1
;
3-
kOH-MMP ) 3 × 109 M-1 s-1, kO -MA ) 0.09 M-1 s-1, kO
3
3-
HiBA ) 5 M-1 s-1; kOH-HiBA ) 3 × 109 M-1 s-1). Since all
compounds reacted slowly with molecular ozone, only the
degradation pathway of MTBE with OH radicals has
been determined, including the formation of primary
degradation products. In experiments performed with
several natural waters, the efficiency of MTBE elimination
and the formation of bromate as disinfection byproduct
have been measured. With a bromide level of 50 µg/L, only
35-50% of MTBE could be eliminated by the AOP O /
3
H O without exceeding the current drinking water standard
2
2
of bromate (10 µg/L). The transient concentrations of
MTBE and its primary degradation products were modeled
using a combination of kinetic parameters (degradation
product distribution and rate constants) together with the
ozone and OH radical concentration and were in good
agreement with the experimental results.
Ozonation can lead to oxidation of micropollutants either
by a direct reaction of the compounds with molecular ozone
or by OH radicals that are produced during the decay of
ozone. The concentration of both oxidants must be known
to define and calibrate an ozonation process with respect to
its oxidation capacity. An experimental approach to deter-
mine the concentration of ozone and OH radicals during
conventional ozonation or the AOP O3/ H2O2 has been
developed in previous studies (14-16). As described by Elovitz
and von Gunten (14), a Rct value can be defined as the ratio
of the exposures of OH radicals and ozone (i.e., concentration
of oxidant integrated over the reaction time). The Rct value
can be calculated from the measurement of the decrease of
Introduction
The widespread occurrence of methyl tert-butyl ether
(MTBE), a major gasoline component, in drinking water
resources such as groundwater and surface water (1) poses
significant problems to the drinking water industry. MTBE
is poorly biodegradable (especially under anaerobic condi-
* Corresponding author telephone: +41-1-8235270; fax: +41-1-
8235210; e-mail: vongunten@eawag.ch.
•
† Present address: Departamento de Ingenieria Quimica, Uni-
versidad de Extremadura, 06071 Badajoz, Spain.
a probe compound, which reacts fast with OH but not with
ozone, and a simultaneous determination of the ozone
9
4 2 5 2 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 21, 2001
10.1021/es010044n CCC: $20.00
2001 Am erican Chem ical Society
Published on Web 09/22/2001

identification of degradation products. For quantitative
analysis, tert-amyl methyl ether (TAME) was selected as the
internal standard, and only five masses were recorded in
selective ion monitoring (SIM) mode, corresponding to the
main mass signals of MTBE and its degradation products (73
for MTBE and TAME and 57, 58, 59, and 74 for TBF, AC, TBA,
and MA, respectively). Aqueous samples (1 µL) without further
pretreatment were injected on-column onto a Stabilwax
(Crossbond Carbowax-PEG) column (60 m × 0.32 mm, 1 µm
film thickness). The oven temperature was kept at 50 °C
during the first 10 min and then increased to 100 °C at 10
°C/ min. After 5 min at 100 °C, the temperature was increased
to 200 °C at 30 °C/ min. Helium was the carrier gas, and the
column head pressure was set at 100 kPa.
The aldehyde intermediates were analyzed by HPLC
(Hewlett-Packard, 1050 series; column: 250 × 4 mm i.d.
Zorbax ODS (5 µm particle size)) after derivatization with
2,4-dinitrophenylhydrazine (2,4-DNPH) following a method
adapted from ref 19. To 5 mL of sample solution, 2 mL of an
aqueous solution containing 800 µM 2,4-DNPH and 500 mM
HCl were added and maintained at room temperature for 2
h before analysis. Separation was performed using gradient
elution at a flow rate of 1 mL/ min. Initial conditions were
45% acetonitrile and 55% water with a linear gradient to 51%
of acetonitrile at 8 min and then increased linearly to 80%
at 15 min. After that, the initial conditions were reached within
2 min, and the system equilibrated for another 8 min. The
sample volume injected varied from 25 to 250 µL depending
on the MTBE concentration. Absorbance was measured at
360 nm. MMP and HiBA were quantified by using trimethyl-
acetaldehyde (TMA) and FA as reference compounds since
these aldehydes are not commercially available.
TABLE 1. Water Quality Parameters
lake water
well water
Zurich
(LZ water)
Murten
Porrentruy
(WP water)
(LM water)
DOC (m g/L)
alkalinity (m M HCO3
pH
1.4
2.7
0.8
-
)
2.5
3.8
5.0
7.8
11.3
28.4
7.8
7.2
brom ide (µg/L)
17.6
16.0
16.2
18.0
am m onia (µg/L of N)
concentration. Once the Rct value is known, the elimination
of a compound (M), which reacts with both oxidants, can be
calculated by second-order kinetics and expressed as a
•
function of Rct, kO , k OH, and the ozone exposure (∫[O3] dt)
3
(16):
[M]
ln
) - ( [O ] dt)(kOHRct + kO
)
(1)
∫
3
(
)
3
[M]0
•
where k OH and kO are the second-order rate constants for
3
the reactions of a micropollutant (M) with •OH and O3,
respectively.
Equation 1 can be used to estimate the elimination of
degradation products as well. In a previous study, the
degradation of atrazine and the formation of its degradation
products could be predicted successfully by measuring the
Rct and the rate constants for the oxidation of atrazine and
its degradation products with ozone and OH radicals (16).
This quantitative kinetic approach, which has also been
applied to asses the oxidation of metal-diethylenetriamine-
pentaacetate (DTPA) complexes during drinking water ozo-
nation (17), will be used to predict the behavior of MTBE and
its primary degradation products during ozonation and the
AOP O3/ H2O2 for various raw waters.
The identification of the aldehyde intermediates was
performed by LC/ MS according to the method described
above. The spectra were obtained on a Platform LC single
quadrupole mass spectrometer equipped with an electro-
spray interface (ESI) from Micromass UK Ltd. (Manchester,
U.K.). The spectra were acquired in negative ion mode over
the mass range of 50-500 Da at 1 scan/ s. The ESI interface
temperature was set to 150 °C, and the nitrogen gas flow was
set to 500 L/ h. The cone voltage was 50 V. Parent ions at m/ z
281 for MMP and 267 for HiBA were found, which corresponds
to (M - H)- ions as observed for other dinitrophenylhy-
drazones.
Materials and Methods
Standards and Reagents. MTBE and its main degradation
products (TBF, TBA, AC, MA, and FA) where obtained from
Aldrich and Sigma, each with a purity above 98%. All
chemicals used for solutions (buffer, eluent, etc.) were reagent
grade and were used without further purification. Stock
solutions were prepared in bidistilled water. Concentrated
ozone solutions were produced by passing O3-containing
oxygen through bidistilled water that was cooled in an ice
bath (18).
Natural Water System s. To evaluate the influence of the
water quality parameters on the degradation of MTBE
together with the formation of degradation products, we
studied the processes in the following Swiss natural waters
(for water quality, see Table 1): (i) raw water of Lake Zurich
collected from a depth of 30 m was obtained from a drinking
water plant in Zurich (LZ water) and is characterized by a
low alkalinity and low DOC; (ii) raw water from Lake Murten
sampled at a depth of 20 m (LM water) is characterized by
a relatively high alkalinity and DOC content; and (iii) raw
well water collected at the drinking water plant in Porrentruy
(WP water). It originates from a karstic system and is
characterized by a high alkalinity and low DOC. All waters
were filtered (0.45 µm cellulose nitrate) within 24 h after
sampling and stored at 4 °C until use.
Characterization of Ozonation and AOP. Dissolved ozone
was analyzed by the Indigo method (20). OH radicals were
indirectly measured by a probe compound, which reacts
slowly with ozone and fast with OH radicals. The selected
probe compound was p-chlorobenzoate (pCBA), which reacts
•
with OH with a second-order rate constant of k ) 5 × 109
M-1 s-1 (21). pCBA was determined by HPLC (column: Merck
Lichrospher 100, RP18 5 µm, 125 × 3 mm i.d.) with an eluent
containing 45% 10 mM H3PO4 and 55% methanol at 1 mL/
min and detected at 234 nm (14). Hydrogen peroxide was
determined by the peroxidase-DPD method (22). Bromide
and bromate were measured by ion chromatography with
postcolumn reaction (23).
Mechanistic Investigations. Oxidation of MTBE and its
degradation products by molecular ozone is too slow to be
used for drinking water treatment. Therefore, AOP O3/ H2O2
experiments with MTBE, TBF, and TBA in bidistilled water
at 20 °C could be used to determine the degradation pathway
with OH radicals. These experiments were carried out by
adding H2O2 (0.1 mM) to the initial solution of these com-
pounds (40-80 µM in all experiments) followed by the
Analytical Methods. MTBE and most of its degradation
products (TBF, TBA, AC, and MA) were analyzed by direct
aqueous injection GC/ MS. Positive electron impact (EI+)
mass spectra were obtained with a Fisons MD800 quadrupole
mass spectrometer interfaced to a Fisons GC8000 gas
chromatograph. Spectra were acquired at an electron energy
of 70 eV. The source temperature was set at 200 °C, and the
interface temperature was set at 250 °C. The mass spectra
were scanned from 25 to 150 (m/ z) at a scan time of 0.5 s for
•
addition of O3. Under these conditions, OH was the pre-
dominant oxidant species. The pH was kept constant at 7
(with 10 mM phosphate buffer). Residual degradation product
concentrations were analyzed after total O3 consumption in
experiments with varying O3 dosages (from 0.1 to 10 mg/ L).
9
VOL. 35, NO. 21, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 4 2 5 3

Experim ents in Natural Waters. Natural water samples
were buffered to pH 7 or 8 by adding 10 mM borate buffer
and adjusting the pH with HCl. MTBE was added at relatively
low initial concentrations (2 µM or 176 µg/ L) to minimize
competition reactions of the oxidants with the natural water
matrix. Furthermore, pCBA was added as a probe compound
(0.5 µM). Bromide was spiked to the samples to reach an
initial concentration of 50 µg/ L. H2O2 was used only in AOP
experiments, with a ratio of 0.34 mg of H2O2/ mg of O3 (0.5
M/ M). After addition of ozone (from 2 to 4 mg/ L), samples
were withdrawn with a dispenser after certain time intervals
(18). The reaction was stopped with indigo for residual ozone
and bromate measurements. For MTBE and pCBA analysis,
residual ozone was quenched by adding a small volume of
a concentrated (1 M) sodium thiosulfate solution.
TABLE 2. Rate Constants for Reaction of MTBE and Its
Degradation Products with Ozone and OH Radicals
compd
kO (M-1 s-1
)
ref
kOH (M-1 s-1
)
ref
3
MTBE
TBF
TBA
MMP
MA
AC
HiBA
FA
0.14
this study 1.9 × 109
this study
this study
21
this study
21
0.78
0.003
5a
0.09
0.032
5a
0.1
5.9
4.2
this study 7.0 × 108
24
6.0 × 108
this study 3.0 × 109
this study 2.3 × 108
a
24
1.3 × 108
1.0 × 109
21
this study 3.0 × 109
this study
21
this study
this study
a
24
TMA
isobutyraldehyde
this study 3.2 × 109
this study 2.8 × 109
a
Estim ate from the values obtained for TMA and isobutyraldehyde.
Determ ination of Rate Constants of MTBE and Its
Degradation Products with Ozone and •OH. The rate
constants of the reaction of ozone with MTBE and its
degradation products were measured by an absolute method
involving the measurement of the decrease of ozone in the
presence of an excess of MTBE to fulfill pseudo-first-order
conditions (24). The temperature was varied from 5 to 20 °C,
and the pH was kept constant at 2 (50 mM phosphate buffer).
Ozone (2 × 10-5 M) was added to buffered solutions of MTBE
and its degradation products (1 × 10-3 M). The ozone
concentration was monitored by measuring its absorbance
at 258 nm (ꢀ ) 3150 M-1 cm-1; 25).
Rate constants of MTBE and some degradation products
with •OH were determined by competition kinetics (26), using
pCBA as a reference compound. OH radicals were produced
by adding different O3 doses (from 0.25 to 5 mg/ L) to solutions
containing MTBE or degradation products (10 µM), pCBA
(10 µM), and H2O2 (0.1 mM) at pH 7 (50 mM phosphate
buffer). The relative decrease of pCBA and MTBE or its
degradation products was measured after complete depletion
of ozone.
kOH values for MTBE (1.9 × 109 M-1 s-1) and TBF (7 × 108 M-1
s-1) fall within the range of published data of 1.6 × 109 (21)
and 3.9 × 109 M-1 s-1 (7) for MTBE and of 4.1 × 108 (27) and
1.2 × 109 M-1 s-1 (7) for TBF.
Oxidation of MTBE and Its Degradation Products by
OH Radicals. To determine the degradation pathway of MTBE
by OH radical attack, experiments were conducted with high
MTBE concentration (1 × 10-4 M) in the presence of H2O2
(2 × 10-4 M) at pH 7 and varying the ozone dosage (from 2
to 10 mg/ L). The identification of the reaction intermediates
was performed by GC/ MS and LC/ MS as described above.
The following degradation products were identified: TBF,
TBA, MMP, AC, MA, HiBA, and FA. These degradation
products have also been identified in previous studies when
other OH radical producing AOPs were applied (UV/ H2O2
(8), sonolysis (12)).
Once the intermediates were identified, the AOP O3/ H2O2
was applied individually to MTBE and its most important
degradation products (TBF and TBA) to quantify the kinetics
of these reactions and to obtain the complete reaction
sequence. These experiments were carried out with an initial
concentration of MTBE or intermediates of 40-80 µM at pH
7 in the presence of H2O2 (1 × 10-4 M) and varying the ozone
dosage (from 0.10 to 10 mg/ L). Under these experimental
conditions, OH radicals are the main oxidant species. The
degradation of MA and AC was not the subject of our study
because of their low reactivity with OH radicals (see Table
2). In addition, the degradation of AC was studied in detail
before (28, 29).
Figure 1 shows the degradation of MTBE and the evolution
of its major degradation products as a function of the ozone
dosage for the AOP system. The primary degradation products
were TBF, TBA, MMP, AC, MA, and FA. A detailed scheme
of MTBE degradation in the UV/ H2O2 system has been
proposed by Stefan et al. (8), including the mineralization of
secondary transformation products. However, some impor-
tant kinetic data including rate constants and percentage of
formation of some intermediates are not reported. From the
available data, it can be estimated that at low ozone doses,
OH radicals react primarily with MTBE while significant
degradation of MTBE transformation products only takes
place after an important fraction of MTBE has been degraded.
Therefore, the percentage of each degradation product
formed from MTBE can be calculated during the initial phase
of a transformation experiment conducted at such conditions
(see Figure 2). The percentages given indicate the fraction
of MTBE oxidized to a particular product. FA has to be
considered separately because it is formed simultaneously
from several degradation products. The total mass balance
has been calculated considering the compounds with high
carbon atom number formed directly from MTBE (sum of
MTBE and degradation products concentration except FA).
The mass balance was close to 100% at very low ozone dose
(see Figure 1) and decreased with increasing ozone dosage
Results and Discussion
Rate Constants of MTBE and Its Degradation Products with
Ozone and OH Radicals. The rate constants of the reaction
of MTBE and of each of its degradation products with both
ozone and OH radicals have to be known to design and
interpret a comprehensive kinetic study on MTBE degrada-
tion in drinking water treatment. Since some of these rate
constants are well-known from the literature (21, 24), only
unknown rate constants or those which are controversial in
the literature were determined in the present study. Our
experimental data are compiled in Table 2 together with
published values. MMP and HiBA were not commercially
available. Therefore, we obtained an estimate of their reaction
rate constants with ozone and OH radicals from the average
of the respective rate constants of structurally related
aldehydes TMA and isobutyraldehyde.
It can be seen that the degradation of MTBE and most of
its intermediates by direct attack of ozone is very slow and
therefore negligible when the AOP O3/ H2O2 is applied. The
oxidation by ozone might be important only for some
aldehyde intermediates.
Rate constants for the reaction of OH radicals (kOH) with
MTBE and some aldehydes (MMP, HiBA, and FA) exceed 109
M-1 s-1, corroborrating previous results of fast degradation
of these compounds by OH radicals. The rate constants of
the other intermediates of MTBE degradation for the reaction
with OH radicals are significantly lower and range from 108
to 7 × 108 M-1 s-1, indicating a potential accumulation of
these compounds during AOPs. FA exhibits a higher reactivity
than AC and MA.
The rate constant for the reactions of MTBE and TBF with
OH radicals have been measured before by competition
kinetics methods (7, 21, 27). Our experimentally determined
9
4 2 5 4 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 21, 2001

butyl side and 60% at the methoxy group. Taking into account
that 42% of MTBE decays to form TBF, this implies that 75%
of MTBE degradation by the OH radical attack on the methoxy
group leads to TBF. This value is somewhat higher than the
66% reported by Stefan et al. (8).
We also investigated the degradation pathway of the
reaction of the main intermediates (TBF and TBA) with OH
radicals as shown in Figure 2. The degradation of TBA leads
to the formation of AC, HiBA, and FA. The fraction of
unidentified primary transformation products was very low
since the mass balance was close to 100%. The same
degradation products were identified during the oxidation
of TBF. However, the mass balance was only 75% after 50%
degradation of TBF, which indicates that degradation prod-
ucts are formed that were not detectable by our analytical
methods. These results disagree with those proposed by
Stefan et al. (8), who reported AC and FA as major primary
intermediates of TBF degradation. In addition, TBA was not
found during the degradation of TBF in accordance with the
relatively slow hydrolysis of TBF to TBA at neutral pH (half-
life of 5 days; 31).
Degradation of MTBE in Natural Waters. In addition to
MTBE, a small amount of a probe compound (pCBA) was
added to the natural waters to quantify the concentration of
OH radicals through the Rct value concept (14). During
conventional ozonation, a biphasic behavior was observed.
An initial phase with high k (first-order rate constant for
ozone decomposition) and Rct values was followed by a
second phase with significantly lower values. For the AOP
O3/ H2O2, only one phase with constant k and Rct values was
noted (see Table 3; only the values of the second phase are
included for conventional ozonation).
FIGURE 1. Concentration profiles of MTBE (left axis) and its
degradation products (right axis) as a function of ozone dose during
the AOP O /H O . Experimental conditions: pH 7, T ) 20 °C, [H O ]
3
2
2
2 2 0
) 1 × 10-4 M, [MTBE]0 ) 40 µM. The data shown represent final
concentrations after complete ozone depletion.
Figure 3 shows the course of MTBE concentration versus
reaction time in experiments performed with LZ water for
different oxidant concentrations and pH values. A relatively
slow MTBE degradation occurred in conventional ozonation
experiments due to the low reactivity of MTBE with molecular
ozone. Addition of hydrogen peroxide (0.34 mg of H2O2/ mg
of O3) accelerated the ozone decomposition into OH radicals
(32), which in turn oxidized MTBE (cf. k and Rct values in
Table 3). The application of an AOP therefore leads to a faster
MTBE degradation. However, even the AOP is not very
effective in removing MTBE from raw waters at oxidant doses
typically applied in drinking water treatment. For instant, a
dose of 2 mg/ L of ozone and a ratio of 0.34 mg of H2O2/ mg
of O3 resulted in a final MTBE conversion of less than 50%
in LZ water. To increase the MTBE degradation, a higher
oxidant concentration must be applied. An O3 dose of 4 mg/ L
yielded 71% of MTBE elimination in LZ water (Figure 3 and
Table 3).
FIGURE 2. MTBE degradation and formation of degradation products.
Percentages given refer to the relative importance of a pathway.
(75% after 90% of MTBE degradation), probably due to the
formation of unidentified intermediates.
Initial OH radical attack on MTBE by H-abstraction can
occur at either the methoxy group or any of the three methyl
groups. The carbon-centered radical generated after addition
of oxygen forming the peroxyl radical decomposes through
bimolecular reactions to different intermediates (30). Reac-
tion at the methoxy group is more likely to happen since
H-abstraction occurs most easily at the position R to the
ether function (30). This reaction pathway generates TBF,
TBA, AC, and FA (8). The second possible pathway (OH radical
attack on the methyl groups) leads to the formation of MMP,
MA, AC, and FA. The contribution to the overall transforma-
tion of MTBE of each pathway can be calculated from the
product distribution data shown in Figure 2. Considering
that MA and AC are generated in equimolecular amounts
from the OH radical attack on a methyl group, the quantity
of AC formed from the OH radical attack on the methoxy
group can be calculated by subtracting the amount of MA
from the total amount of AC. Thus, around 44% AC originate
from OH radical attack on the three methyl groups, and 56%
are generated from OH radical attack on the methoxy group.
This result confirms that the presence of oxygen in the
molecule facilitates the electrophilic attack of OH radicals
on the methoxy group (56% of the total process). This ratio
agrees well with the data from Stefan et al. (8), who found
40% MTBE decay as a result of OH radical attack at the tert-
An increasing pH leads to a faster ozone decomposition
according to the higher values of k and Rct obtained at pH
8 (Table 3). This faster ozone decomposition into OH radicals
leads to a faster MTBE degradation but not to a higher degree
of MTBE elimination (Figure 3).
The influence of various water quality parameters on the
O3 decay and OH radicals formation can be evaluated by
comparing the values of k and Rct given in Table 3 for each
natural water in experiments referring to similar experimental
conditions. An increase of NOM concentration accelerates
ozone decomposition into OH radicals as indicated by the
increasing values of k and Rct at higher DOC concentrations.
This is due to the promoting effect of NOM, which results
in an acceleration of the radical-type chain reactions for ozone
decomposition. The influence of the water quality parameters
on k and Rct values is more pronounced in conventional
ozonation because in the AOP O3/ H2O2 the initiation of O3
-
decomposition by HO2 is an important step and NOM
promotion becomes less relevant in the whole process. The
influence of the water quality parameters on MTBE degra-
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TABLE 3. MTBE Removal, Bromate Formation, Ozone Decomposition Rate Constant, and R Values for Natural Waters Treatment
ct
pH 7
pH 8
[O ] ) 2 mg/L
3
[O ] ) 4 mg/L
3
[O ] ) 2 mg/L
3
ozone
ozone/H O
ozone/H O
ozone
ozone/H O
2 2
2
2
2
2
LZ Water (DOC: 1.4 mg/L; Alkalinity: 2.5 mM; Scavenging Capacity: 5.6 × 104 s-1
)
% elim MTBE
39
46
71
35
36
4.2
brom atea (µg/L)
15.1
8.8
16.8
12.5
k (s-1
Rct
)
1.1 × 10-4
5.3 × 10-3
6.8 × 10-3
4.7 × 10-4
1.2 × 10-2
7.4 × 10-8
6.4 × 10-10
4.4 × 10-8
6.1 × 10-8
2.8 × 10-9
WP Water (DOC: 0.8 mg/L; Alkalinity: 5 mM; Scavenging Capacity: 6.2 × 104 s-1
)
% elim MTBE
28
37
65
22
brom atea (µg/L)
20.7
12.5
33.5
12.5
k (s-1
Rct
)
6.8 × 10-5
3.8 × 10-3
5.3 × 10-3
2 × 10-4
2.2 × 10-10
1.9 × 10-8
3.6 × 10-8
4 × 10-10
LM Water (DOC: 2.7 mg/L; Alkalinity: 3.8 mM; Scavenging Capacity: 10 × 104 s-1
)
% elim MTBE
33
34
46
brom atea (µg/L)
12
5.6
8.7
k (s-1
Rct
)
7.7 × 10-4
8.3 × 10-3
1 × 10-2
4.4 × 10-8
4 × 10-9
4.2 × 10-8
a
Initial brom ide concentration in the experim ents: 50 µg/L.
FIGURE 3. Influence of the oxidant concentration and pH on MTBE
degradation. Experimental conditions: LZ water, T ) 12 °C, ratio
of H O /O ) 0.34 w/w, [MTBE]0 ) 1.8 µM. Symbols represent
measured data, and lines represent model calculations. Final
reaction time for conventional ozonation experiments were 6 and
2 h at pH 7 and pH 8, respectively.
FIGURE 4. Influence of water source on MTBE degradation.
Experimental conditions: T ) 12 °C, pH 7, [O ] ) 4 mg/L, ratio of
3 0
H O /O ) 0.34 w/w, [MTBE]0 ) 1.8 µM. Symbols represent measured
2
2
3
2
2
3
data, and lines represent model calculations.
bromate in many countries (10 µg/ L) (33, 34), a multi-
objective optimization of the process is often required.
Bromate formation occurs through a complicated mechanism
involving both molecular ozone reactions and reactions with
OH radicals (35). In hydrogen-peroxide-based AOPs, the
reduction of hypobromous acid (HOBr) by hydrogen peroxide
is a crucial process (36) that reduces bromate formation (37,
38). However, in the AOP O3/ H2O2, bromate formation cannot
be completely suppressed even at very high levels of hydrogen
peroxide (37). In Table 3, MTBE oxidation and concomitant
bromate formation are shown for three different waters under
various oxidation conditions. All waters were spiked with
bromide to yield an initial concentration of 50 µg/ L. The
same basic features occurred in all three of the waters and
are discussed in detail below for LZ water. Since MTBE only
reacts with OH radicals at significant rates, it can be taken
as a measure for the OH radical exposure. For an ozone dose
of 2 mg/ L, addition of H2O2 increased the MTBE elimination
at pH 7 only slightly as compared to the conventional
ozonation test, indicating that the OH radical exposure
remained almost unchanged. However, in the AOP O3/ H2O2,
the half-life of molecular ozone is much smaller and therefore
the ozone exposure decreases. Thus the overall oxidation
dation is depicted in Figure 4 for experiments carried out
under conditions yielding maximum MTBE degradation (pH
7, ozone dose of 4 mg/ L, and presence of H2O2). The yield
of MTBE degradation was highest in LZ water and lowest in
LM water due to the differences in the overall scavenging
capacities of the waters rather than due to the different ozone
decomposition rates into OH radicals. The overall OH radical
scavenging capacities are defined as follows: kOH-DOC[DOC]
+ kOH-HCO [HCO3-] + kOH-CO [CO32-]; kOH-DOC ) 2.5 × 104
-
2-
3
3
6
mg-1 s-1L (25), kOH-HCO ) 8.5 × 10 M-1 s-1, kOH-CO ) 3.9
-
2-
3
3
× 108 M-1 s-1 (21). At pH 7, the values for the scavenging
capacities are 5.6 × 104, 6.2 × 104, and 1.0 × 105 s-1 for LZ
water, WP water, and LM water, respectively. These values
can explain the trend of the MTBE degradation yield observed
in Figure 4 and Table 3 except those for the conventional
ozonation of WP water and LM water (28% and 33%
elimination, respectively).
Oxidation of MTBE and Brom ate Form ation. In bromide-
containing waters, the ozone-induced oxidation of mi-
cropollutants is always accompanied by the formation of
bromate. Because of the low drinking water standard for
9
4 2 5 6 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 35, NO. 21, 2001

capacity (OH radicals and ozone) decreases in the system.
In addition, the presence of H2O2 leads to a partial reduction
of HOBr, which is an important intermediate in bromate
formation. These two factors reduced the bromate formation
by approximately a factor of 2, from 15.1 to 8.8 µg/ L.
Kinetic Modeling of MTBE Degradation by the AOP O3/
H2O2. A kinetic model was applied to predict the evolution
of MTBE and its degradation products in natural waters by
conventional ozonation and the AOP O3/ H2O2 and compared
to experimental results. For this, in addition to the rate
constants and the degradation product distribution, the
concentration of both oxidants ozone and OH radicals were
determined (k and Rct in Table 3). The contribution of several
degradation pathways must be considered to estimate the
transient formation of the degradation products. For that
purpose, the overall rate constants of MTBE and of each of
its degradation products with both oxidants have been split
up according to the partial contribution of each pathway
given in Figure 2. With the determined oxidant concentrations
and the split rate constants, the concentrations of MTBE
and its degradation products have been calculated from eq
1 with the help of the computer program ACUCHEM (39).
At higher ozone doses (4 mg/ L), an increased bromate
formation is found because of the increased overall oxidation
capacity in the system. An increase of the ozone dose leads
to an enhanced elimination of MTBE (by almost a factor of
2). This shows that the OH radical exposure is increased
significantly with the higher ozone dose. In addition, the
ozone exposure also increases. Both factors lead to an
increased bromate formation by about a factor of 2 (from 8.8
to 16.8 µg/ L). Therefore, MTBE oxidation in the LZ water is
restricted by bromate formation at pH 7. An oxidation of
about 50% can be achieved without exceeding the bromate
standard of 10 µg/ L when the AOP O3/ H2O2 is applied with
an initial bromide concentration of 50 µg/ L.
The data shown in Figures 3 and 4 allow comparisons of
the model calculations of MTBE degradation (lines) with
measured data (symbols) for various experimental conditions
and types of natural waters. It can be seen that MTBE
oxidation is well predictable. Figure 5 shows measured data
(symbols) and model calculations (lines) for MTBE removal
and formation of its degradation products during treatment
of LZ water (Figure 5a), WP water (Figure 5b), and LM water
(Figure 5c) at the respective experimental conditions (for
details, see figure captions). Experiments performed with
natural waters resulted in measurable concentrations of
MTBE, TBF, TBA, AC, and MA. Because of the low initial
concentration of MTBE (2 µM) and the interference of
thiosulfate in the analytical method, some aldehyde inter-
mediates could not be quantified. FA could only be measured
in the last reaction sample (after complete ozone depletion),
and the obtained values were very similar to those predicted
by model calculations. The concentrations of MMP (also a
primary degradation product) as predicted by the kinetic
model are also included in Figure 5 (dotted line), although
we cannot compare these values to measured data. The
dashed line in Figure 5 represents the mass balance including
measured concentrations of the primary degradation prod-
ucts MTBE, TBF, TBA, AC, and MA and the modeled
concentrations of MMP. Some control samples of natural
waters before and after ozonation have been analyzed, and
no MTBE degradation products were found (data not shown),
which indicates that the measured mass balance is in fact
due to MTBE degradation. The experimental mass balance
was well above 90% of the initial MTBE concentration in all
cases, indicating that the formation of the secondary
degradation products is not relevant until MTBE degradation
exceeds 50%. Note that measured concentrations of AC were
often slightly higher than the model predictions (mainly in
conventional ozonation experiments), which can be at-
tributed to the formation of AC from oxidation of NOM or
small interferences in the background signal of the GC/ MS
from the natural water. The otherwise good agreement
between the model calculations and the experimental results
validates the results of our study.
An increase in pH from 7 to 8 leads to a decrease of the
bromate formation for both conventional ozonation and the
AOP O3/ H2O2. To compare both pH values, however, the
elimination of MTBE has to be considered. In both cases,
less MTBE is eliminated at the higher pH, which is the result
of a lower OH radical exposure. In addition, the experiments
show that the ozone exposure is lower (k increases) at the
higher pH. Another reason for a reduced bromate formation
is the enhanced reduction of HOBr by H2O2 at the higher pH
value. The superposition of all three factors results in a lower
bromate formation (36). Only when normalized to the same
ozone exposure bromate formation increases with increasing
pH (35).
The two other waters (WP, LM) follow basically the same
trend (see Table 3) as discussed for LZ water. However, the
degree of bromate formation varied significantly. The
comparison of LZ water and WP water for the AOP O3/ H2O2
with an ozone dose of 4 mg/ L shows that the MTBE oxidation
was similar, whereas the bromate formation was almost twice
in WP water. In the DOC-poor WP water, the half-life of
ozone was substantially higher, which lead to a higher ozone
exposure. In addition, due to the high carbonate content,
carbonate radicals contributed to enhanced bromate forma-
tion (35). Therefore, in WP water only about 35% of MTBE
can be oxidized by the AOP O3/ H2O2 without exceeding the
bromate standard with the initial bromide concentration used
in these experiments (50 µg/ L). In LM water, about 50% of
MTBE can be oxidized without exceeding the bromate
standard, but a higher ozone dose is required due to the high
overall scavenging capacity of this water. In fact, the same
degree of MTBE oxidation can be achieved for LM water
with an ozone dose of 4 mg/ L (AOP O3/ H2O2) for LZ water
with an ozone dose of 2 mg/ L (AOP O3/ H2O2). The bromate
formation in both cases is identical at about 8.8 µg/ L, and
the ozone exposure is similar as well (1.4 × 10-4 M min-1).
The factor of 2 in ozone dosage needed to obtain similar
results is reflected in a 2-fold greater scavenging capacity for
LM water as compared to LZ water.
In conclusion, MTBE oxidation by conventional ozonation
and the AOP O3/ H2O2 has to be carefully optimized with
regard to bromate formation. To achieve a similar degree of
MTBE oxidation, the overall scavenging capacity has to be
considered and the ozone dose adapted accordingly. Once
a certain degree of MTBE oxidation has been set, the bromate
formation is a function of the resulting ozone exposure. The
oxidation of MTBE is an “unfortunate” case because of its
relatively low reactivity toward OH radicals. For other
compounds with higher rate constants with OH radical (e.g.,
aromatic compounds), better removal could be achieved,
and the optimization pollutant elimination/ bromate forma-
tion would be easier.
Overall, the combination of experimental kinetic data with
modeling enables to predict not only the time course of MTBE
removal but also the concomitant formation of its degradation
products during conventional ozonation and AOP treatment
of natural waters. In addition, our results allow a quantitative
prediction of the formation of primary degradation products
based on the degree of MTBE elimination. This is shown in
Figure 6 where the production of degradation products is
correlated to the MTBE elimination (molar basis). The data
shown in Figure 6 include both conventional ozonation and
the AOP O3/ H2O2 for the different water types investigated
in the present study. The slopes of the straight lines are 0.35
(TBF and FA), 0.23 (AC), 0.11 (TBA), and 0.07 mol/ mol (MA).
9
VOL. 35, NO. 21, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 4 2 5 7

FIGURE 6. Formation of degradation products as a function of MTBE
elimination. Intermediate and final amounts obtained in experiments
shown in Table 3 are included.
due to its higher reactivity toward ozone and OH radicals,
which leads to its further degradation (Figure 6 inset).
Our experiments at the lower level of MTBE (2 µM) showed
a good mass balance for product formation even though
they were extrapolated from high level MTBE experiments.
A similar behavior can be expected for concentrations <2
µM that are more representative for polluted environmental
systems because MTBE is not a major sink for OH radicals
under these conditions. Therefore, the present kinetic
approach can also be applied in water works with much
lower MTBE concentrations (i.e., 0.5-20 µg/ L). This is
possible since our concept is based on the relative MTBE
elimination and formation of degradation products and is
independent of absolute concentrations.
Acknowledgments
We acknowledge G. Amy for reviewing the manuscript, R.
Schoenenberger for LC/ MS analysis, and E. Salhi and L.
Zwank for technical support. J.L.A. is grateful to the Junta de
Extremadura for providing him necessary funds to ac-
complish this research.
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