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Ozone oxidation of pharmaceuticals, endocrine disruptorsand pesticides during drinking water treatment

R. Broseus a,*, S. Vincent a, K. Aboulfadl b, A. Daneshvar c, S. Sauve b,B. Barbeau a, M. Prevost a

a NSERC Industrial Chair on Drinking Water, Civil, Geological and Mining Engineering Department, Ecole Polytechnique de Montreal,

CP 6079, Succ. Centre-Ville, Montreal, QC, Canada H3C 3A7b Department of Chemistry, Universite de Montreal, P.O. Box 6128, Succ. Centre-Ville, Montreal, QC, Canada H3C 3J7c Department of Aquatic Sciences and Assessment, Swedish University of Agricultural Sciences, Box 7050, 750 07 Uppsala, Sweden

a r t i c l e i n f o

Article history:

Received 5 May 2009

Received in revised form

22 July 2009

Accepted 25 July 2009

Published online 6 August 2009

Keywords:

Drinking water

Ozone

Caffeine

Pharmaceuticals

Endocrine disruptors

Pesticides

* Corresponding author. Tel.: þ1 514 340 471E-mail address: romain.broseus@polymtl

0043-1354/$ – see front matter ª 2009 Elsevidoi:10.1016/j.watres.2009.07.031

a b s t r a c t

This study investigates the oxidation of pharmaceuticals, endocrine disrupting compounds

and pesticides during ozonation applied in drinking water treatment. In the first step,

second-order rate constants for the reactions of selected compounds with molecular ozone

ðkO3 Þ were determined in bench-scale experiments at pH 8.10: caffeine (650� 22 M�1 s�1),

progesterone (601� 9 M�1 s�1), medroxyprogesterone (558� 9 M�1 s�1), norethindrone

(2215� 76 M�1 s�1) and levonorgestrel (1427� 62 M�1 s�1). Compared to phenolic estrogens

(estrone, 17b-estradiol, estriol and 17a-ethinylestradiol), the selected progestogen endo-

crine disruptors reacted far slower with ozone. In the second part of the study, bench-scale

experiments were conducted with surface waters spiked with 16 target compounds to

assess their oxidative removal using ozone and determine if bench-scale results would

accurately predict full-scale removal data. Overall, the data provided evidence that ozone is

effective for removing trace organic contaminants from water with ozone doses typically

applied in drinking water treatment. Ozonation removed over 80% of caffeine, pharma-

ceuticals and endocrine disruptors within the CT value of about 2 mg min L�1. As expected,

pesticides were found to be the most recalcitrant compounds to oxidize. Caffeine can be

used as an indicator compound to gauge the efficacy of ozone treatment.

ª 2009 Elsevier Ltd. All rights reserved.

1. Introduction through conventional wastewater treatment (Halling-Sor-

Pharmaceuticals and personal care products (PPCPs), endo-

crine disrupting compounds (EDCs) and pesticides are groups

of organic micropollutants which are routinely detected in

surface waters (Heberer et al., 1997; Benotti et al., 2009a;

Mompelat et al., 2009) and even in drinking waters (Stackel-

berg et al., 2004; Jones et al., 2005; Snyder et al., 2007a). A

significant fraction of the PPCPs and EDCs released into the

aquatic environment is attributed to their incomplete removal

1 (#2194); fax: þ1 514 340.ca (R. Broseus).er Ltd. All rights reserved

ensen et al., 1998; Daughton and Ternes, 1999; Joss et al., 2004;

Clara et al., 2005; Segura et al., 2007). The most representative

pharmaceutical compounds detected in urban wastewaters

are anti-inflammatory drugs, anticonvulsants, antibiotics and

lipid regulators (Suarez et al., 2008). Commonly reported

hormones are the natural steroid estrogens estrone (E1),

17b-estradiol (E2), estriol (E3) and the synthetic contraceptive

17a-ethinylestradiol (EE2). Despite the low concentration

levels of EDCs, their occurrence has been linked to undesirable

5918.

.

Table 1 – Target compounds.

Compound Class/use MDL(ng L�1)

MW(g mol�1)

pKa

Caffeine Stimulant 9 194.2 10.4a

Trimethoprim Anti-infective 9 290.3 7.12a

Carbamazepine Anticonvulsant 2 236.3 0.37a

Naproxen Analgesic 12 230.3 4.15a

Gemfibrozil Anti-

cholesterol

24 250.3 4.42a

Estrone Estrogen 10 270.4 10.4a

Estriol Estrogen 50 272.4 10.4a

Estradiol Estrogen 3 288.4 10.4a

17a-Ethinylestradiol Synthetic

estrogen

7 296.4 10.4a

Progesterone Progestrogen 3 314.5 NA

Medroxyprogesterone Synthetic

progestrogen

2 344.5 NA

Norethindrone Synthetic

progestogen

7 298.4 NA

Levonorgestrel Synthetic

progestogen

5 312.4 NA

Cyanazine Herbicide 4 240.7 1.1b

Deethylatrazine

(DEA)

Metabolite of

atrazine

5 187.6 1.4b

Deisopropylatrazine

(DIA)

Metabolite of

atrazine

17 173.6 1.5b

NA¼Not applicable; MDL¼method detection limit; MW¼molecular weight.

a Reference: Snyder et al. (2007b).

b Reference: Viglino et al. (2008b).

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 74708

effects and toxicological impacts in fish species (Desbrow

et al., 1998; Kramer et al., 1998; Snyder et al., 2001).

Some pesticides are also endocrine disrupting compounds

(Ikehata et al., 2008). Possible sources of pesticide contami-

nation in drinking water sources include agricultural and

urban runoffs, direct application to control insects and vege-

tation, domestic usage and leaching from pesticide wastes

(Ikehata and El-Din, 2005). Atrazine is among the most

frequently detected pesticides in surface and drinking water

at concentrations ranging from the ng L�1 to mg L�1 level

(Hua et al., 2006). Due to their persistence in aquatic envi-

ronment and potential adverse health effects, contamination

of surface water and groundwater with pesticides has long

been recognized as an important issue in many countries.

Due to the large number of drugs consumed throughout

the world, their complex chemical structure and potential

hazard to human health via the consumption of drinking

water, there is a growing interest in understanding the fate of

PPCPs and EDCs during drinking water treatment (Ternes

et al., 2002; Benotti et al., 2009b). Treatment options have been

suggested for the removal of these trace organic contami-

nants. Conventional drinking water treatment processes

(coagulation/flocculation, filtration, chlorination) have been

shown to be fairly ineffective in completely removing atrazine

and selected PPCPs and EDCs (Ternes et al., 2002). Recent

studies have shown that reverse osmosis, nanofiltration and

adsorption on granular and powdered activated carbon were

effective for removing those contaminants (Westerhoff et al.,

2005; Snyder et al., 2007b). However, several compounds were

detectable in membrane permeate and carbon effluent, and

removal efficiency using activated carbon was largely depen-

dent on water quality (Snyder et al., 2007b). In addition,

physical separation processes produce wastes which have to

be properly disposed of (Ikehata et al., 2008).

Chemical oxidation using ozone has been proved to be an

effective treatment process for a wide spectrum of organic

micropollutants during bench-, pilot- and full-scale experi-

ments in both wastewater and drinking water (Ternes et al.,

2002, 2003; Huber et al., 2005; Westerhoff et al., 2005; Ikehata

et al., 2006; Snyder et al., 2006; Hua et al., 2006; Vieno et al.,

2007; Gagnon et al., 2008). Huber et al. (2005) showed that

municipal wastewater effluents spiked with 11 selected PPCPs

and treated with ozone in a pilot-scale were oxidized as much

as 90–99% at ozone doses ranging from 2 to 5 mg O3 L�1. Ozone

(2.5 mg L�1) was also found to be highly effective in 20 drinking

water treatment plants (DWTPs) from diverse locations across

the United States, while chlorine (2.5 mg L�1) and UV

(40 mJ cm�2) were less efficient (Snyder, 2008). The pH,

temperature, water characteristics and ozone dose have an

impact on the degradation of organic micropollutants (Yar-

geau and Leclair, 2008), although a comprehensive under-

standing of these effects is still needed.

Due to its high oxidation potential, ozone treatment is

widely used in drinking water treatment for disinfection, color

removal, taste and odor control, decrease of disinfection by-

products formation, biodegradability increase and also for the

successful degradation of many organic contaminants. Ozone

reacts with organic contaminants through both a direct

reaction with molecular ozone or through indirect reactions

with free radicals (including the hydroxyl radical OH�)

produced by the decomposition of ozone. The rate of OH�

formation depends on the water matrix, especially its pH,

alkalinity, type and content of natural organic matter (von

Gunten, 2003). Molecular ozone reacts selectively with

unsaturated bonds, aromatic systems and amino groups

whereas the reaction with OH� radicals is a faster and unse-

lective process.

To assess the removal efficiency of ozonation, it is neces-

sary to determine the rate constants for the reaction of

micropollutants with ozone and OH� radicals. Rate constants

for direct ðkO3 Þ and indirect (kOH) ozone reactions with some

PPCPs, EDCs and pesticides have been measured and pub-

lished in the literature (von Gunten, 2003; Huber et al., 2003).

Those values can be used to predict the effectiveness of ozone

oxidation. Previous studies generally focused on natural

hormones (E1–E3) and the most commonly detected phar-

maceuticals (Huber et al., 2003; Deborde et al., 2005). However,

only limited information is available about ozone efficiency

against caffeine, progesterone and synthetic progestogen

steroids (medroxyprogesterone, norethindrone, levonorges-

trel), which are used in various hormone therapies (Viglino

et al., 2008a). In addition, most ozone studies have been con-

ducted at bench- or pilot-scale. It was of interest to evaluate if

bench-scale results were in agreement with data gathered in

full-scale ozone treatment processes in the greater region of

Montreal (Canada). In that sense, this paper is a unique

contribution to the various works focused on the ozone

oxidation to improve the quality of drinking water (Ternes

et al., 2003; Huber et al., 2003; Snyder et al., 2006).

Table 2 – Filtered water quality for DWTP A and DWTP B.

Parameter Unit DWTP A DWTP B

pH 8.24 6.62

Alkalinity mg CaCO3 L�1 80 9

Dissolved organic

carbon (DOC)

mg C L�1 1.99 2.85

UV absorbance (254 nm) cm�1 0.027 0.047

Turbidity NTU 0.112 0.058

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 7 4709

2. Objectives of study

In the research project presented hereafter, four specific

objectives were sought:

1. To assess the potential of ozone, and specifically molecular

ozone, to oxidize caffeine, progesterone and synthetic

progestogen steroids (medroxyprogesterone, norethin-

drone and levonorgestrel). The laboratory tests involved the

determination of the second-order reaction rate constants

for the reaction of the selected compounds with ozone.

Bench-scale assays were carried out in pure aqueous

solution and in natural waters in order to assess the impact

of water quality on ozone efficacy;

2. To verify the ozone oxidation of 16 organic contaminants

(pharmaceuticals, steroids and pesticides) spiked at bench-

scale in filtered water samples;

3. To determine if those bench-scale results could be used to

estimate full-scale drinking water treatment plant (DWTP)

removal data;

4. To evaluate if one target contaminant could be used as

a surrogate to evaluate ozone efficacy against organic trace

contaminants.

3. Materials and methods

3.1. Standards and reagents

Table 1 shows the target compounds including caffeine,

4 pharmaceuticals (trimethoprim, carbamazepine, naproxen

and gemfibrozil), 8 natural and synthetic steroids (estrone,

17b-estradiol, estriol, 17a-ethinylestradiol, progesterone,

medroxyprogesterone, norethindrone and levonorgestrel) and

3 pesticides (cyanazine, DIA and DEA). All standards were

obtained from Sigma–Aldrich Canada (Oakville, ON, Canada).

Internal standards included [13C3]-caffeine, [13C3]-atrazine and

[13C2]-estradiol and were supplied by ACP Chemical Inc.

(Montreal, QC, Canada). All solvents (trace analysis grade),

0.1% formic acid water and LC-grade water were purchased

from Baker (QC, Canada). Individual stock solutions were

prepared by dissolving accurately-weighed samples in ultra-

pure water in order to obtain final concentrations ranging

from 2 to 10 mg L�1. These solutions were placed in an ultra-

sonic bath at 35 �C for different times, depending on the

compounds, and then stored in the dark at 4 �C. These solu-

tions were filtered over 0.22 mm filters. Methanol was not used

for dissolution of the target compounds in order to prevent the

impact of the solvent on ozone decay rates and OH� concen-

tration. Ultrapure water (18 MU cm) was produced with a Milli-

Q (Millipore, USA) apparatus. Phosphate buffer and tert-

butanol (final concentrations: 0.2 M and 50 mM, respectively)

were prepared by dissolution of the commercial compounds

in water.

3.2. Natural water samples

Bench-scale experiments were performed using two natural

waters that differed in alkalinity and dissolved organic carbon

concentration (DOC). Filtered water samples (before the

ozonation process) were collected from two municipal DWTPs

in the province of Quebec, the first one (DWTP A) with

moderate alkalinity (80 mg CaCO3 L�1) and low DOC

(1.99 mg C L�1), and the second one (DWTP B) with low alka-

linity (9 mg CaCO3 L�1) and moderate DOC (2.85 mg C L�1).

Those waters were sampled in 10 L polypropylene carboys

that had been thoroughly washed and rinsed successively

with distilled and ultrapure waters. Natural water samples

were filtered (0.45 mm polyethersulfone) and kept at 4 �C prior

to the ozone experiments. Conventional water quality anal-

yses, presented in Table 2, were conducted for natural

samples filtered on 0.45 mm following the standard methods.

Those surface waters contained a few detectable PPCPs, EDCs

and pesticides, but in all such cases, the detected concentra-

tions were much lower than the concentration spiked during

the bench-scale experiments.

3.3. Analytical methods

3.3.1. Quantification of PPCPs, EDCs and pesticidesMethods used for the detection and quantification of PPCPs,

EDCs and pesticides were described previously (Viglino et al.,

2008a,b). Briefly, the target compounds were analyzed by

automated on-line solid phase extraction (SPE) coupled with

liquid chromatography (LC) and tandem mass spectrometry

(MS/MS). Two ionisation sources, electrospray ionisation (ESI)

and atmospheric pressure photoionisation (APPI), were used

for the simultaneous detection of PPCPs and pesticides, and

EDCs, respectively.

3.3.2. Ozone analysisAt bench-scale, ozone stock solution concentration and ozone

residual in water were determined according to the standard

colorimetric method 4500-O3 (APHA, 1998) using indigo tri-

sulfonate (3600nm¼ 20 000 M�1 cm�1). Samples were analyzed

at 600 nm with a spectrophotometer Varian (Cary 100,

Victoria, Australia) in a 1-cm or 2-cm quartz cell.

3.4. Bench-scale experiments

Bench-scale ozone experiments were performed at 20� 1 �C

with buffered ultrapure water and filtered water samples

collected from both municipal water treatment plants. For rate

constant determinations, water samples were spiked with

target compounds to achieve the following concentrations:

2754� 246 ng L�1 (caffeine), 2074� 310 ng L�1 (progesterone),

1801� 248 ng L�1 (medroxyprogesterone), 2847� 162 ng L�1

-5.00

-4.00

-3.00

-2.00

-1.00

0.000.00 1.00 2.00 3.00 4.00 5.00 6.00

CT (mg.min.L-1) CT (mg.min.L

-1)

CT (mg.min.L-1) CT (mg.min.L

-1)

ln

(C

/C

0)

-5.00

-4.00

-3.00

-2.00

-1.00

0.000.00 1.00 2.00 3.00 4.00 5.00 6.00

-6.00

-5.00

-4.00

-3.00

-2.00

-4.00

-3.00

-2.00

-1.00 -1.00

0.000.00 2.00 4.00 6.00 8.00 10.00

ln

(C

/C

0)

ln

(C

/C

0)

ln

(C

/C

0)

-5.00

0.000.00 0.50 1.00 1.50 2.00

a b

c d

Fig. 1 – Evolution of normalized caffeine (a), progesterone (b), medroxyprogesterone (c) and norethindrone (d) concentration

with CT value (pH 8.10). Symbols: , buffered ultrapure water; - natural water.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 74710

(norethindrone) and 2344� 180 ng L�1 (levonorgestrel). The

verification of the ozone oxidation of the 16 organic contami-

nants was conducted using initial concentrations between 100

and 200 ng L�1. During kinetic runs (objective #1), tert-butanol

was added to quench OH� radicals generated by ozone

decomposition. The pH of ultrapure water was adjusted with

a phosphate buffer while the pH of natural waters was left

unbuffered. Stock ozone solutions (50–60 mg O3 L�1) were

preparedby diffusing gaseous ozone produced with an oxygen-

fed generator (Ozone services, BC, CA) through a water-jack-

eted flask containing ultrapure water chilled at 4 �C. Applied

ozone doses (0–3 mg L�1) were administered by injecting an

aliquot of this ozone stock solution via a syringe into a true

batch continuously stirred 1-L glass reactor containing the

Table 3 – Second-order rate constants (±standarddeviation) for the reactions of molecular ozone with theselected compounds (pH [ 8.10, T [ 20 ± 1 8C).

Compound kO3 (M�1 s�1)

BUW NW

Caffeine 650� 22 573� 6

Progesterone 601� 9 630� 25

Medroxyprogesterone 558� 9 510� 9

Norethindrone 2215� 76 2563� 60

Levonorgestrel 1427� 62 –

BUW¼ buffered ultrapure water; NW¼ natural filtered water.

water sample and equipped with a floating Teflon lid to prevent

ozone degassing. After ozone addition, aliquots of 4 mL were

withdrawn from the test water at regular time intervals for

ozone residual analysis. For any target compound analysis,

aliquots of 40 mL were withdrawn and transferred into vials

containing ascorbic acid (5 g L�1) to quench the residual ozone,

stop the oxidation reaction and prevent microbial degradation.

Westerhoff et al. (2005) showed that ascorbic acid was suitable

to quench oxidant residuals during sampling and did not affect

the stability of some EDCs and PPCPs. Target compounds’

concentrations were analyzed for various contact times (from

0 to 15 min). Bench-scale ozone tests provided information

regarding ozone demand and ozone decomposition. The

instantaneous (initial) ozone demand was measured as the

difference between the ozone dose and the ozone residual after

w10 s. The results were analyzed using Statistica software

version 7.1 (StatSoft Inc., 2006). Unless otherwise mentioned,

standard deviations were used to characterize uncertainty in

kinetic constants evaluation.

3.5. Determination of rate constants for the reactionwith molecular ozone

The value of the rate constant with molecular ozone was

determined for each compound at pH 8.10 in buffered ultra-

pure water and at ambient pH in natural water (DWTP A), in

the presence of tert-butanol as hydroxyl radical scavenger.

The kinetics of ozone reactions with organic and inorganic

compounds is typically second order, i.e. first order both with

ozone and the contaminant concentrations (von Gunten,

Table 4 – Structure of the investigated compounds in thekinetic assays (Viglino et al., 2008a,b).

Compound Structure

Caffeine

N

NOCH3

O

NN

NCH3

H3C

Progesterone

O

O

H

H

H

Medroxyprogesterone

O

O

OH

HH

H

Norethindrone

Levonorgestrel

OH

O

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 7 4711

2003). The degradation of caffeine, progesterone, medrox-

yprogesterone, norethindrone and levonorgestrel with

molecular ozone can be described by:

�d½TC�dt

¼ kO3½O3�½TC� (1)

where TC¼ target compound and kO3¼ reaction rate constant

with molecular ozone. The rate constant is obtained from the

integration of Eq. (1):

ln

�½TC�t½TC�

�¼ �kO3

Z t

½O3�dt (2)

0 0

where ![O3]dt is the time-integrated ozone concentration also

termed ozone exposure. The value of the second-order rate

constant can be found from the gradient of a plot of ln removal

of the target compound against the time-integrated ozone

concentration.

The time-integrated ozone concentration defines an ozone

exposure CT (concentration� contact time). In this study, CT

values (mg min L�1) were performed using the integrated CT

concept (Barbeau et al., 2005), for which the effective CT at

time t (min) is equal to the area under the decay curve at that

time. CT values were calculated using the ozone concentra-

tion profiles (Eq. (3)) assuming a simple first-order decay:

CTeffective ¼Z

CðtÞdt ¼ Co

k0�1� exp

��k0,t

��(3)

where C¼ ozone residual (mg L�1); Co¼ initial ozone residual

(mg L�1) determined from the exponential fit of the relation

between the ozone residual and the time (min); k0 ¼ ozone

first-order decay constant (min�1). CT values in mg min L�1

were converted in M s for subsequent comparison of reaction

kinetics (M�1 s�1) with data from the literature.

3.6. Full-scale evaluation

Water samples were collected before and after the ozone

process at both drinking water facilities. Full-scale disinfec-

tion process efficiency was calculated using the CT10 concept

(USEPA, 1989). Ozone residuals were quenched using ascorbic

acid. While all target compounds were analyzed in the labo-

ratory, only some of them were detectable in natural samples.

4. Results and discussion

4.1. Kinetic rate constants

Second-order rate constants for the reaction of caffeine,

progesterone, medroxyprogesterone, norethindrone and lev-

onorgestrel were determined in buffered ultrapure water and

natural water during direct ozonation (i.e. with tert-butanol) to

assess the oxidation of the compounds. The ozonation

experiments were duplicated and undertaken at pH 8.10. Fig. 1

presents the kinetic data from the experiments for caffeine,

progesterone, medroxyprogesterone and norethindrone

(R2> 0.97). The kinetic reaction rate constants are compiled in

Table 3.

Overall, reaction rate constants varied from 558� 9 M�1 s�1

(medroxyprogesterone) to 2215� 76 M�1 s�1 (norethindrone)

in ultrapure water buffered to pH 8.10. The constants calcu-

lated here for the progestogen steroids are far lower than

those reported for steroid phenolic hormones

(kO3 � 106 M�1 s�1 at the same pH) (Huber et al., 2003; Deborde

et al., 2005). Progesterone and medroxyprogesterone exhibited

quite similar reaction rate constants, which may be due to

their similar chemical structure (Table 4). The removal of

progesterone by ozone was studied by Barron et al. (2006). The

rate constant, which was independent of pH in the presence of

-2.00

-3.00

-4.00

-5.00

-6.00

-1.00

0.000.00 1.00 2.00 3.00 4.00 5.00 6.00 7.00

CT (mg.min.L-1

)

ln

(C

/C

0)

Fig. 2 – Evolution of normalized caffeine and progesterone

concentration with CT value in buffered ultrapure water

(pH 6.30). Symbols: , caffeine; - progesterone.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 74712

tert-butanol (in the pH range of 2.0–8.0), was evaluated to be

equal to 480� 30 M�1 s�1 and varied from 444� 11 M�1 s�1 to

521� 15 M�1 s�1. In this study, the rate constant in presence of

tert-butanol was slightly higher at 601� 9 M�1 s�1. To our

knowledge, no data have yet been published in the literature

for caffeine, norethindrone and levonorgestrel. However,

caffeine has been reported to exhibit an antioxidant activity

and to scavenge highly reactive free radicals, including

hydroxyl radicals (Dalmazio et al., 2005). Norethindrone and

levonorgestrel contain ethynyl groups which may exhibit

higher reactivity towards ozone. This could explain why

higher kinetic constants were observed for their reaction with

ozone as compared with caffeine, progesterone or medrox-

yprogesterone. In the case of the direct reaction with molec-

ular ozone, the kinetic rate constants are influenced by the

properties of the target substances. From the rate constants

determined in this work, the half-life of the target contami-

nants varied from 0.25 min (norethindrone) to 1.0 min

(medroxyprogesterone) while using an ozone concentration of

about 1 mg L�1 (21 mM).

Progesterone, medroxyprogesterone, norethindrone and

levonorgestrel do not present any acid or basic character.

Moreover, the pKa of caffeine (10.4) will favour the presence of

the protonated form for conditions currently encountered in

0.00

0.25

0.50

0.75

1.00

1.25

1.50

1.75

2.00

2.25

2.50

Time (min)

Ozo

ne

re

sid

ua

l (m

g.L

-1

)

1 mg/L O3

2 mg/L O3

2 mg/L O3+ tert-butanol

3 mg/L O3

0 2 4 6 8 10

a

Fig. 3 – Ozone decay for DWTP A (a) and DWTP

drinking water production processes (5.0–9.0 pH range).

Therefore, it is not anticipated that the rate constant for those

compounds will be strongly pH-dependent when OH� radicals

formation is inhibited (by adding tert-butanol, for example).

This was verified with similar bench-scale experiments for

caffeine and progesterone by adding those compounds in

ultrapure water buffered to pH 6.30. The kinetic data are

shown in Fig. 2 (R2> 0.92). The kinetic rate constants were

evaluated to be equal to 695� 42 M�1 s�1 and 564� 8 M�1 s�1 at

pH 6.30 for caffeine and progesterone, respectively. The

kinetic rate constants were equal to 650� 22 M�1 s�1 and

601� 9 M�1 s�1 at pH 8.10, respectively, confirming the pH

value did not largely influence the rate constants.

In this study, the impact of the water matrix on the reac-

tion rate constant was unclear and varied according to the

contaminant of interest. In natural waters spiked with

a radical scavenger (DWTP A), the reaction rate constant

decreased 11.8% for caffeine and 8.7% for medroxyprogester-

one, increased 13.5% for norethindrone and there was no

significant impact for progesterone (Table 3). Globally, kinetic

rate constants were in the same order of magnitude. Reaction

rates determined in ultrapure water could therefore be used to

approximately predict the oxidation of the four compounds

dissolved in natural waters and exposed to direct molecular

ozonation. This result was also confirmed by Huber et al.

(2003, 2005), who studied the reaction of selected pharma-

ceuticals with ozone in natural waters.

4.2. Oxidation with ozone of spiked organiccontaminants in natural waters

Dissolved ozone decay curves for the two plants filtered

waters are presented in Fig. 3. Overall, four ozone oxidation

experiments were performed per DWTP: in filtered waters at 1,

2 and 3 mg O3 L�1, as well as in filtered water at 2 mg O3 L�1

with tert-butanol (50 mM). Addition of tert-butanol, acting as

an OH� radical inhibitor, allowed us to evaluate the natural

impact of OH� promotion by comparison of the ozone molec-

ular pathway for two identical ozone doses. In both natural

waters, the ozone residual experienced a classic initial period

of high decay rate followed by a second period of slower decay.

Ozone residuals decreased faster in DWTP B filtered waters

(k0 ¼ 0.12–0.84 min�1) compared to DWTP A filtered waters

Time (min)

1 mg/L O3

2 mg/L O3

2 mg/L O3+ tert-butanol

3 mg/L O3

0.00

0.25

0.50

0.75

1.00

1.25

1.50

1.75

2.00

2.25

2.50

0 2 4 6 8 10

Ozo

ne

re

sid

ua

l (m

g.L

-1

)

b

B (b) natural filtered waters (T [ 20 ± 1 8C).

Table 5 – Rate constants for the reactions of target compounds with molecular ozone and OH� radicals.

Compounds kO3 (M�1 s�1) T (�C)a kOH (M�1 s�1) References

Caffeine 650� 22 20� 1.0 5.9–6.9� 109 This study; Kesavan and Powers, 1985

Trimethoprim 2.7� 105 20� 0.5 6.9 (�0.2)� 109 Dodd et al., 2006

Carbamazepine w3� 105 20 8.8 (�1.2)� 109 Huber et al., 2003

7.8 (�1.3)� 104 25 2.05� 109 Andreozzi et al., 2002; Vogna

et al., 2004

Naproxen w2� 105 20 9.6 (�0.5)� 109 Huber et al. 2005

Estrone 6.2� 103–2.1� 107 w20 1.1� 109–7� 1010 Nakonechny et al., 2008

1.53� 105–4.24� 109 20� 2 – Deborde et al., 2005

17b-Estradiol 106 20 1.41� 1010 Huber et al., 2003; Rosenfeldt

and Linden, 2004

2.21� 105–3.69� 109 20� 2 – Deborde et al., 2005

Estriol 1.01� 105–3.89� 109 20� 2 – Deborde et al., 2005

17a-Ethinylestradiol 3� 106 20 9.8 (�1.2)� 109 Huber et al., 2003

1.83� 105–3.65� 109 20� 2 1.08 (�0.23)� 1010 Deborde et al., 2005; Rosenfeldt

and Linden, 2004

Progesterone 480� 30 18� 1 – Barron et al. 2006

601� 9 20� 1 – This study

Medroxyprogesterone 558� 9 20� 1 – This study

Norethindrone 2215� 76 20� 1 – This study

Levonorgestrel 1427� 62 20� 1 – This study

Cyanazine 7.34–61.8 – 1.9� 109 Ikehata and El-Din, 2005; Camel

and Bermond, 1998

DIA 7.5 – 2.1� 109 Beltran et al., 2000

DEA 0.2 – 2� 109 Beltran et al., 2000

Atrazine 6–7.9 – 2.4–3.0� 109 Westerhoff et al., 2005

a T (�C) refers to kO3 studies.

wa

te

rr

es

ea

rc

h4

3(2

00

9)

47

07

–4

71

74

71

3

Table 6 – Maximum concentrations of target compounds(ng LL1) during sampling campaigns at DWTP A andDWTP B.

DWTP A DWTP B

Before O3 After O3 Before O3 After O3

Caffeine 214 60 267 54

Trimethoprim 19 <9 <9 <9

Carbamazepine 8 4 10 4

Naproxen <12 <12 28 34

Gemfibrozil <24 <24 <24 <24

Estradiol 3 <3 5 <3

Cyanazine 7 5 <4 <4

DIA 237 29 <17 <17

DEA 57 57 10 10

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 74714

(k0 ¼ 0.14–0.31 min�1). DWTP B water was more reactive, as

shown by a higher initial demand (35–53%) than in DWTP

A water (28–41%). This observation is consistent with their

respective DOC concentration (1.99 vs 2.85 mg C L�1). Using

a 2 mg L�1 ozone dose with and without tert-butanol only

induced a significant difference in ozone residual profile in

DWTP B water. This difference might be due to the significant

alkalinity in DWTP A waters (80 mg CaCO3 L�1) which already

acted as OH� scavengers. Bicarbonate/carbonate are known to

inhibit ozone decomposition due to their ability to scavenge

OH� radicals (Staehelin and Hoigne, 1985).

Analysis showed that gemfibrozil and naproxen were not

detected in spiked ozonated water samples (data not shown).

The estrogenic steroids (estrone, 17b-estradiol, estriol, and

17a-ethinylestradiol) which contain phenolic moieties (Huber

et al., 2003), as well as carbamazepine and trimethoprim, were

not detected or detected at trace levels close to the detection

limit. In the DWTP B natural water, progesterone, medroxy-

progesterone, norethindrone and levonorgestrel were

removed below the method detection limit (MDL). In the

DWTP A water, the removal of norethindrone, medrox-

yprogesterone and levonorgestrel was �95%, �57% and �88%

at 1.03 mg min L�1, respectively (progesterone was not

spiked). These data demonstrate the potential of ozone for the

0.0

10.0

20.0

30.0

40.0

50.0

60.0

70.0

80.0

90.0

100.0

0.00 1.00 2.00 3.00 4.00 5.00

CT (mg.min.L-1

)

Ca

ffe

in

e re

mo

va

l (%

)

a

Fig. 4 – (a) Comparison of caffeine removal observed at lab-scal

full-scale; dotted lines: 90th percentile confidence intervals; (b)

observed at full-scale. Symbols: C caffeine/lab-scale (kO3 [ 650

caffeine/full-scale; outliers.

oxidation of selected pharmaceuticals and endocrine dis-

ruptors in drinking water treatment facilities. Caffeine

removal was more limited under the range of CT tested and

varied from 23% to 80% at 0.27–1.85 mg min L�1 in DWTP

A filtered waters. Caffeine was also detected in ozonated

natural water samples where the activity of OH� radicals was

strongest (DWTP B). Overall, ozonation removed over 80% of

caffeine, PPCPs and EDCs within the CT value of about

2 mg min L�1 for both source waters at 20� 1 �C. It is not

determined at this time if these performances will be signifi-

cantly impacted by water temperature.

As expected, DEA and DIA which are by-products of atra-

zine chemical oxidation, and the herbicide cyanazine were

found to be the most recalcitrant compounds to oxidize.

Those compounds do not contain aromatic moieties but an

electron-withdrawing group (chloro-substituted double-bond)

that leads to this low reactivity towards ozone (von Gunten,

2003). At the maximum calculated CT value (13.2 mg min L�1),

the removal of pesticides was greater in DWTP B than in

DWTP A waters: 38% vs 4%, 38% vs 12%, and 65% vs 46% for

DEA, DIA, and cyanazine, respectively. This result is linked to

the higher OH� radicals exposure in DWTP B filtered waters.

Viglino et al. (2008b) showed that caffeine, carbamazepine

and 5 pesticides including cyanazine, DIA and DEA were

detected at the water intake of DWTP A. In this study, rapid

removals were observed for some spiked selected compounds

whereas pesticides were less susceptible to ozone. The results

obtained are consistent with the kinetics of the reaction of

ozone previously determined in bench-scale experiments in

ultrapure water for various pH conditions (Table 5). While kO3

values vary between 0.2 (DEA) and >109 M�1 s�1 (estrogenic

steroids), hydroxyl radical rate constants only differ by one

order of magnitude due to the lower selectivity of OH� radicals.

The rate constants for the hydroxyl radicals are typically around

109 M�1 s�1 (Westerhoff et al., 2005). Based on the concentration

of target compounds spiked in natural waters and their actual

MDL, other experiments performed in 2008 showed that ozone

doses typically applied in drinking water treatment

(2 mg O3 L�1) were able to achieve removals higher than 98.8%

for pharmaceuticals (trimethoprim, carbamazepine, naproxen

and gemfibrozil) and 96.0% for estrogenic steroids (estrone, 17b-

0.00 1.00 2.00 3.00 4.00 5.00

CT (mg.min.L-1

)

0.0

10.0

20.0

30.0

40.0

50.0

60.0

70.0

80.0

90.0

100.0

Ca

ffe

in

e re

mo

va

l (%

)

b

e and full-scale. Symbols: C caffeine/lab-scale; caffeine/

comparison of caffeine removal predicted at lab-scale and

ML1 sL1); B caffeine/lab-scale (kO3 [ 573 ML1 sL1);

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 7 4715

estradiol, estriol and 17a-ethinylestradiol), respectively. The

rate constants of reaction of molecular ozone with these

compounds are greater than 7.8� 104 M�1 s�1. According to von

Gunten (2003), compounds with kO3 > 104 M�1 s�1 react quickly

with molecular ozone at doses typically applied in drinking

water treatment.

Most of the target contaminants were previously studied in

natural waters during bench-, pilot- and full-scale experi-

ments with doses similar to those used in this study. Adams

et al. (2002) demonstrated rapid conversion of trimethoprim

with an ozone dose of 0.3 mg L�1. Naproxen was not detected

at ozone doses equal to 0.4 mg O3/mg total organic carbon

(Vieno et al., 2007). Trimethoprim, carbamazepine, naproxen

and gemfibrozil removal was higher than 98% in ozonated

water samples using an ozone dose of 2.5 mg L�1 and a 2-min

contact time (Snyder et al., 2006). High reactivity of carba-

mazepine towards ozone was also reported elsewhere (Ternes

et al., 2002; Vieno et al., 2007) while estrogenic steroids were

completely removed when ozone was applied (Huber et al.,

2003, 2005; Ternes et al., 2003; Snyder et al., 2006). Kinetic rate

constant of atrazine, cyanazine and DIA for the reaction with

ozone is quite similar (Table 5). Atrazine exhibited <50%

removal using an ozone dose of 2.5 mg L�1 (Snyder et al., 2006)

but the removal of pesticides can be maximized using

advanced oxidation processes (UV/O3, O3/H2O2), as shown

during our experiments with the formation of OH� radicals.

For the same conditions, caffeine and progesterone removal

varied from 91% to >99% (Snyder et al., 2006).

4.3. Prediction of full-scale removal data

Water samples were collected before and after the ozone

disinfection process at both municipal DWTPs to determine if

previous bench-scale results would accurately predict full-scale

removal data. The waters were sampled 11 times from April

2007 to November 2007. The concentrations of pharmaceuticals,

steroids and pesticides as well as the parameters describing the

water characteristics at the time of sampling were determined

in these samples. Maximum concentrations obtained varied

depending on the compound and the type of natural water

(Table 6). The dominating compound was the psycho-stimulant

caffeine. Gemfibrozil was never detected in both natural waters

contrary to caffeine. In DWTP B, progesterone was detected only

once before ozonation but was not detected after ozonation.

Pesticides were always detected in the DWTP A water. One

explanation previously suggested was a greater cumulative

agricultural load in the St-Lawrence River upstream of DWTP A

(Viglino et al., 2008b).

Full-scale disinfection process efficiency was calculated

using the CT10 concept developed by the USEPA (1989). This

allowed, for a specific target compound, to correlate the

removal at bench-scale and full-scale resulting from ozone

oxidation. However, the comparison was limited for DWTP

B results since only caffeine and pesticides had concentra-

tions higher than the MDL at bench-scale. Some of the

caffeine removal data had to be rejected because in some

cases the full-scale concentrations were below the MDL and

no correlation could be established for pesticides. Only

caffeine removal at bench-scale was remarkably similar to

removal observed in full-scale DWTPs except for three outliers

(Fig. 4a). Moreover, the caffeine reaction rate constants

determined in the first step adequately predicted the ozone

oxidation of caffeine observed at full-scale (Fig. 4b). The

kinetic rate constants for the reaction of caffeine with

molecular ozone are lower than those of most organic

contaminants and caffeine was also found to be persistent

during drinking water treatment (Daneshvar et al., 2008). If

one wishes to analyze for caffeine, a simplified and shorter

procedure can be used for the chromatography and MS

determinations, thus allowing for a higher throughput of

environmental samples. Caffeine is highly soluble and has

negligible volatility (Dalmazio et al., 2005) and can easily be

spiked in laboratory or bench-scale studies. Caffeine was also

present in higher concentrations and at a higher frequency

than most pharmaceuticals and EDCs, consistent with the

observations made in source waters by Daneshvar et al. (2008).

Our data suggest the use of caffeine as an indicator compound

to monitor and predict the efficacy of ozone treatment at all

experimental scales and for a wide range of compounds

including oxidation-recalcitrant hormones. Caffeine has been

used with some success as a chemical marker for domestic

wastewater due to its wide consumption (Buerge et al., 2003).

Finally, public communication of analytical monitoring of

caffeine results carries less public stigmata than those of

minute quantities of various pharmaceuticals and EDCs.

Estrone, trimethoprim, DEET, atenolol, atrazine and mepro-

bamate can serve as oxidation (Cl2 and ozonation) indicators

(Benotti et al., 2009a), but their measurement is more complex

than a simple determination of caffeine concentration.

5. Conclusion

Oxidation kinetics of caffeine, progesterone, medroxy-proges-

terone, norethindrone and levonorgestrel with molecular ozone

ðkO3 Þ were investigated in ultrapure and natural waters. The

second-order rate constants for direct molecular ozone reaction

were determined as 650� 22 M�1 s�1, 601� 9 M�1 s�1,

558� 9 M� 1s�1, 2215� 76 M�1 s�1 and 1427� 62 M�1 s�1 in

ultrapure water buffered to pH 8.10, respectively. It was found

that a variation in pH did not largely influence the rate constant

for caffeine and progesterone. The impact of the water matrix

was also minimal but varied according to the contaminant of

interest. When the 16 target compounds (pharmaceuticals,

steroids and pesticides) were spiked in natural waters at bench-

scale, theremoval resultswere in linewithotherpublished work

on this topic. Overall, ozonation removed over 80% of caffeine,

PPCPs and EDCs (13 compounds) within the CT value of about

2 mg min L�1 for both source waters. As expected, pesticides

were found to be the most challenging compounds to oxidize,

with removals less than 65% (cyanazine) at 13.2 mg min L�1.

It was found that caffeine removal at bench-scale was similar to

removal observed in full-scale DWTPs. It is proposed that

caffeine could be used as an indicator compound to predict full-

scale removal of ozonation processes for the oxidation of a wide

range of EDCs and pharmaceuticals. The oxidation by-products

of PPCPs, EDCs and pesticides were not investigated in this

study. However, they should be identified in order to gather

information about the potential impact associated with their

eco- and toxicological effects.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 4 7 0 7 – 4 7 1 74716

Acknowledgments

The authors would like to acknowledge the technical staff

from the CREDEAU and the NSERC Industrial Chair on

Drinking Water for their support in the laboratory work and

their assistance with sample collection. The NSERC Industrial

Chair on Drinking Water of Ecole Polytechnique and its part-

ners (City of Montreal, City of Laval and John Meunier Inc.) as

well as the Canadian Foundation for Innovation (CFI) provided

funding for this project.

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