ARTICLE IN PRESS
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Water Research 39 (2005) 1628–1638
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Kinetics of reactions between chlorine and the cyanobacterialtoxins microcystins
Juan L. Aceroa,�, Eva Rodrigueza, Jussi Meriluotob
aDepartamento de Ingenieria Quimica y Energetica, Faculdad de Ciencias, Universidad de Extremadura, Avda. Elvas s/n,
06071 Badajoz, SpainbDepartment of Biochemistry and Pharmacy, Abo Akademi University, P.O. Box 66, 20521 Turku, Finland
Received 14 September 2004; received in revised form 25 January 2005; accepted 26 January 2005
Available online 11 March 2005
Abstract
Blooms of cyanobacteria can give rise to the production of toxins which contaminate drinking water sources. Among
the oxidants and disinfectants typically applied in waterworks, chlorine has been found to be effective for the
degradation of microcystins. In the present study, unknown second-order rate constants for the reactions of
microcystin-LR (MC-LR), -RR and -YR with chlorine were determined over a wide pH range. It was found that an
increase of pH has a negative effect on the microcystin degradation rate.
Apparent second-order rate constant for the chlorination of MC-LR at 20 1C varied from 475M�1 s�1 at pH 4.8 to
9.8M�1 s�1 at pH 8.8. From these apparent second-order rate constants, rate constants for the reactions of MC-LR
with hypochlorous acid (HOCl) and hypochlorite (ClO�) were evaluated. Half-life times ranged from minutes at pH 6
to 1 h at pH 8 for a constant residual chlorine concentration of 1.0–0.5mg l�1, typical of oxidation pre-treatment and
final disinfection. Similar reactivity with chlorine was found for MC-RR and MC-YR. Therefore, chlorination is a
feasible option for microcystin degradation during oxidation and disinfection processes, and can be applied in drinking
water treatment in case of cyanobacterial toxin risk if the pH is kept below 8.
Published by Elsevier Ltd.
Keywords: Cyanobacteria; Microcystins; Chlorine; Chloramine; Rate constants; Drinking water chlorination
1. Introduction
Toxic cyanobacteria (blue-green algae) have been
reported on all continents (Chorus and Bartram, 1999).
Cyanobacteria are considered an important water
quality problem and a hazard in regard to their potential
to contaminate water intended for drinking, recreational
and agricultural purposes (Lawton and Codd, 1991;
Chorus and Bartram, 1999). Many blooms are highly
acutely toxic and have been responsible for the death of
e front matter Published by Elsevier Ltd.
atres.2005.01.022
ing author. Tel./fax: +34 924289385.
ess: [email protected] (J.L. Acero).
both wild and domestic animals, as well as human illness
and death resulting from exposure to blue-green algal
toxins (Soong et al., 1992; Turner et al., 1990; Frazier
et al., 1998; Jochimsen et al., 1998). Therefore, drinking
water authorities world-wide are dealing with the
challenge of treating contaminated water.
A number of genera of cyanobacteria can produce
lethal toxins, which have been classified in two diverse
groups, namely hepatotoxic cyclic peptides and neuro-
or cytotoxic alkaloids (Chorus and Bartram, 1999). In
addition, all cyanobacteria contain pyrogenic and
nausea-causing lipopolysaccharides in their cell walls
(Weckesser et al., 1979). Within the cyclic peptide group,
ARTICLE IN PRESSJ.L. Acero et al. / Water Research 39 (2005) 1628–1638 1629
microcystins are highly acutely hepatotoxic and they
also show tumor promotion activity through inhibition
of protein phosphatases 1 and 2A (PP1 and PP2A)
(Yoshizawa et al., 1990). Microcystins, the most wide-
spread cyanobacterial hepatotoxins, are produced
mainly by cyanobacteria belonging to the genera
Microcystis, Anabaena, Planktothrix and Nostoc (Spoof,
2004). At least 76 different microcystin analogues have
been found in natural blooms and laboratory cultures of
cyanobacteria. The most common variant is microcystin
LR (MC-LR), which has been identified together with
other microtoxin variants such as MC-LA, MC-RR and
MC-YR in natural water samples (Yoo et al., 1995;
Falconer et al., 1999; Spoof et al., 2003). The chemical
structure of all microcystins is very similar, they are
cyclic peptides consisting of seven amino acid residues
with two main variable residues. All microcystins
contain a variant of the Adda residue (3-amino-9-
methoxy-2,4,8-trimethyl-10-phenyldeca-4(E),6(E)-die-
noic acid), the structural unit considered crucial for the
toxicity (An and Carmicael, 1994). MC-LR is the most
toxic microcystin, with an LD50 value of 50mg/kg(Dawson, 1998). Because of the potent toxicity of
MC-LR, the World Health Organization (WHO) has
set a provisional guideline value of 1.0 mg l�1 for MC-LRin drinking water (WHO, 1998). This WHO guideline
value is used as a basis in some countries such as Spain
to set a drinking water standard for total microcystins of
1.0mg l�1 (BOE, 2003).MC-LR can be eliminated from natural waters
with the oxidants and disinfectants typically applied
in waterworks. Oxidation investigations using ozone
(Rositano et al., 1998, 2001), chlorine (Nicholson et al.,
1994; Nicholson and Rositano, 1997) and UV photolysis
in presence of TiO2 (Lawton et al., 1999) have shown
that MC-LR is readily oxidised to non-toxic degrada-
tion products under appropriate conditions. Chlorami-
nation, however, did not have any significant effect on
MC-LR (Nicholson et al., 1994). Chlorine was selected
in the present work since it is used worldwide in large
quantities for preoxidation treatment and final disinfec-
tion (Bull et al., 1995). Although the ozonation process
is faster (Rositano et al., 1998), Nicholson et al. (1994)
have reported that microcystins can be destroyed by
chlorine as long as sufficient dose is present, the
oxidation being a pH dependent process. Chlorine
residual of 0.5mg l�1 after 30min contact time was
effective at pHo8, whereas inadequate chlorine dosesand higher pH caused negative results. Tsuji et al. (1997)
have identified dihydroxymicrocystin as the main
degradation product, proposing a reaction mechanism
in which HOCl likely attacks to the Adda group of MC-
LR to form a chloronium ion at the conjugated diene of
Adda, followed by hydrolysis to form dihydroxymicro-
cystin. The chlorination of microcystin-LR is not
believed to lead to the formation of noxious products
(Tsuji et al., 1997). Moreover, degradation of micro-
cystin by chlorination of drinking water did not result in
an increased incidence of cancer in mice genetically
predisposed to the effect of genotoxic carcinogens
(Senogles-Derham et al., 2003).
Although the chlorination of microcystins has been
extensively examined, there has not been a thorough
investigation to determine chlorination rate constants
useful to predict degradation efficiencies during full-scale
treatment processes. Therefore, the objective of the
present study was the calculation of second-order rate
constants for the reaction of chlorine with MC-LR and
other variants such as MC-RR and MC-YR. The
influence of some operating variables such as pH and
temperature on chlorination rate constants was also
established. In addition, a reaction mechanism was
proposed allowing the calculation of individual rate
constants for the reaction of MC-LR with chlorine species
(hypochlorous acid and hypochlorite). Based on these
chlorination rate constants, the required oxidant exposure
(CT value) to reduce MC-LR concentration to 1mg l�1
was determined under several treatment conditions.
2. Experimental section
All reagents were of the purest available quality.
Solutions of analytical reagents, chlorine, and phosphate
buffers were prepared with ultra-pure water produced
from a MilliQ (Millipore) water purification system.
Stock solutions of chlorine were prepared by diluting a
commercial solution of sodium hypochlorite (4% active
chlorine, Aldrich) and standardized spectrophotometri-
cally in presence of excess of iodide to form triiodine
(e at 351 nm ¼ 25,700M�1 cm�1) (Bichsel and von
Gunten, 1999). Monochloramine solution was prepared
daily by adding HOCl to ammonium nitrate at pH 8
with a molar ratio of 1:1. Such solution was standar-
dized by the ABTS method (Pinkernell et al., 2000).
2.1. Toxin materials
Purified toxins: MC-LR and MC-RR to be used in the
oxidation experiments were purified from Anabaena
strain 90 (originally from the culture collection of
Prof. Kaarina Sivonen, University of Helsinki) accord-
ing to Meriluoto et al. (2000). In summary, the
cyanobacteria were extracted with 50% methanol. The
extracts were diluted with water and concentrated by
C18 (Bond Elut, Varian) flash chromatography. The
extracts were further fractionated by silica gel chroma-
tography using a 47� 25mm I.D. Si60 column (Merck)
and a mobile phase of 0.1% trifluoroacetic acid-90%
methanol in water. The toxins were then subjected to
preparative HPLC using a Hichrom Nucleosil 100–10
C18 column, 150� 21.2mm I.D., and a mobile phase of
ARTICLE IN PRESSJ.L. Acero et al. / Water Research 39 (2005) 1628–16381630
27% acetonitrile-73% ammonium acetate, 1 g l�1, in
water. Finally the toxins were desalted on C18 solid-
phase extraction columns. For calibration purposes on
the analytical HPLC instrument (method E in Meriluoto
et al., 2000) a further purified grade of MC-LR was
used. The additional purification was done using a
Supelco Discovery C18 column, 250� 10mm I.D., and a
gradient mobile phase of 0.05% trifluoroacetic acid in
water and acetonitrile.
Toxin extracts: Freeze-dried cyanobacterial material
from Sulejow reservoir, Poland, collected in 2002, was
extracted with 50% methanol, cleaned-up with C18 flash
chromatography, filtered, divided into aliquots and
dried in vials with argon gas. The microcystins in the
extract were determined by analytical HPLC and found
to be MC-LR, MC-RR and MC-YR.
2.2. Analytical methods
The chlorine and chloramine concentration was
analyzed by the ABTS method (Pinkernell et al.,
2000). Spectrophotometric measurements were per-
formed with a Unican Helios b spectrophotometer. pHmeasurements were carried out with a Crison electrode
and a Crison GLP-22 pH-meter, which was calibrated
with standard buffer solutions (Crison).
Concentrations of microcystins was determined by
HPLC (Waters Alliance 2695 Separation Module and
Waters 996 Photodiode Array Detector). A preconcen-
tration step was not needed with the MC concentration
of 1.5–3.5mg l�1 used in the experiments. The column
was a C18 end-capped Merck Purospher STAR RP-18e,
3mm particles, LiChroCART 55� 4mm I.D. and the
injection volume 10ml. Gradient elution (flow rate
1mlmin�1) of acetonitrile with 0.05% trifluoracetic acid
(A) and water with 0.05% trifluoracetic acid (B) was
used by varying the volume percentage of A from 25%
to 70% over 5min. The photodiode Array (PDA)
detector was set in the range 200–300 nm, being
238 nm the wavelength selected for quantification. The
concentrations of MC-RR and MC-YR were estimated
using the same detector response as for MC-LR.
2.3. Rate constants of chlorine and monochloramine with
MC-LR
Chlorination experiments were conducted under
pseudo-first-order conditions where chlorine was at least
20 folds in excess. Ultra-pure water produced from a
MilliQ system was used to prepare the reaction samples.
The temperature was varied in the range 10–30 1C and
the pH between 4 and 9 with 10mM phosphate buffer.
In a typical experiment, a 5ml volume of buffered MC-
LR solution (1.5–2.5mg l�1, 1.5–2.5 mM) was preparedin a batch reactor (serum vial of 5.5ml), which was
capped with PTFE-faced silica septum and located in a
thermostatic bath. The experiments were started after
addition of an aliquot of the chlorine stock solution into
the reactor while stirring (for 5 s). Thus the initial
chlorine concentration ranged from 49 to 223 mM. Atfixed time intervals, 0.3ml of sample was rapidly
transferred with a syringe into a HPLC vial containing
2ml of thiosulfate (0.1M) to stop the reaction. Theconcentration of residual MC-LR was analyzed directly
by HPLC. All experiments were duplicated.
Similar experiments were performed with monochlor-
amine to determine the rate constant of its reaction with
MC-LR at pH 8 and 20 1C. In this case, and due to the
low reactivity of monochloramine, the influence of the
pH and the temperature was not investigated. The initial
concentration of oxidant ranged from 400 to 720 mM,high enough to observe an appreciable degradation of
MC-LR within several hours. The initial concentration
of toxin was 2.4mg l�1.
2.4. Rate constants of chlorine with MC-RR and MC-YR
The determination of the rate constants for the
reactions of HOCl with MC-RR and MC-YR was
performed by using a competitive kinetics model. Direct
measurements were not performed to minimize HPLC
analysis errors (MC-RR and MC-YR concentrations
had to be estimated as MC-LR equivalent). To control
the accuracy of analysis MC-LR was present in these
experiments. MC-LR, which rate constant was already
determined, was used as the reference compound. In
addition, MC-YR was not available as pure toxin. Then,
couples of target and reference compounds were selected
to determine unknown rate constants.
The experiments with MC-RR were carried out in
5.5ml serum vials at 20 1C and pH ranging from 6 to 8
(10mM phosphate buffer). The initial solutions contained
equal concentration of MC-LR and -RR (2.3mM). Afteraddition of chlorine (100mM) while vigorously stirring (for5 s), samples were withdrawn at fixed time intervals and
rapidly transferred into HPLC vials containing tiosulphate
to quench the reaction. Remaining concentrations of
target and reference compounds were analyzed by HPLC.
Rate constants of MC-YR chlorination could be
measured in a toxin extract from Sujelow Reservoir
(Poland), in which MC-YR was present together with
MC-LR and MC-RR. A similar experiment to those
described for MC-RR was performed with this toxin
extract. Again, MC-LR was selected as reference
compound and MC-YR as target compound.
3. Results and discussion
3.1. Kinetics of the reaction of chlorine with MC-LR
The rate constant for the reaction of HOCl with
MC-LR was investigated under pseudo-first-order
ARTICLE IN PRESS
0 50 100 150 200-1.2
-0.9
-0.6
-0.3
0.0
ln([
MC
-LR
]/[M
C-L
R] 0
)
time, s
Fig. 1. Pseudo-first-order kinetic plot for the chlorination of
MC-LR at 20 1C and pH 7.2. The solid line is a linear least-
squares regression of the data. Initial concentrations:
[HOCl]0 ¼ 49.2mM, [MC-LR]0 ¼ 2.35mM (’); [HOCl]0 ¼
95.7mM, [MC-LR]0 ¼ 2.30mM (&); [HOCl]0 ¼ 142.5mM, [MC-LR]0 ¼ 2.35mM (K); [HOCl]0 ¼ 222.8mM, [MC-LR]0 ¼
2.50mM (J); [HOCl]0 ¼ 96.2mM, [MC-LR]0 ¼ 1.49mM (m);
[HOCl]0 ¼ 94.6mM, [MC-LR]0 ¼ 1.91mM (B).
Table 1
Chlorination experiments performed at pH 7.2 and 20 1C
Expt. [MC-LR]0 (mM) [Cl2]t0 (mM) k0 (� 103 s�1)
1 2.35 49.2 3.49
2 2.30 95.7 7.21
3 2.35 142.5 9.60
4 2.50 222.8 16.2
5 1.91 94.6 7.50
6 1.49 96.2 7.27
J.L. Acero et al. / Water Research 39 (2005) 1628–1638 1631
conditions ([HOCl]b[MC-LR]) in a batch system
between pH 4 and 9. Previous investigations on
chlorination of phenolic compounds over a wide pH
range have demonstrated that the reaction between
chlorine and phenols could be described by second-order
kinetics, first-order in active oxidant (HOCl+OCl�) and
first-order in phenol (Rebenne et al., 1996; Gallard and
von Gunten, 2002a). Based on these studies and taking
into account the presence of a conjugated double bond
in the MC-LR molecule, it can be assumed that the
reaction of chlorine with MC-LR follows second-order
kinetics:
�d½MC-LR�
dt¼ kapp½MC-LR�½Cl2�t ¼ k0
½MC-LR� (1)
where kapp is the apparent second-order rate constant
(pH dependent as will be discussed later), [Cl2]t is the
total concentration of chlorine species (HOCl+ClO�)
and k0 is the pseudo-first-order rate constant. The total
chlorine concentration can be considered constant
during the experiment, thus being k0 ¼ kapp [Cl2]t0. The
integration of Eq. (1) leads to:
ln½MC-LR�
½MC-LR�0¼ �k0t (2)
Therefore, if the assumption of first-order kinetics
with respect to MC-LR is correct, the representation of
the first term of Eq. (2) versus time must lead to a
straight line for each experiment. Fig. 1 presents the
results from experiments performed at a range of
MC-LR and chlorine concentrations and at pH 7.2. As
can be observed, points lie satisfactorily around straight
lines, whose slopes are the pseudo-first order rate
constants. After linear regression analysis (r240.99),the calculated values of k0 are summarized in Table 1
together with the initial conditions in these experiments.
The rate of MC-LR degradation increases with higher
initial chlorine dose and is not dependent on the initial
MC-LR concentration. Fig. 2 illustrates a plot of k0
values versus the initial concentration of chlorine. A
linear plot could be drawn (r240.99), which confirmsthat the rate of MC-LR degradation is also first-order
with respect to chlorine concentration. The slope from
linear analysis of data in Fig. 2 yields an apparent
second-order rate constant at pH 7.2 of 72.0M�1 s�1.
The rate of MC-LR degradation, therefore, is first-order
in the concentrations of chlorine and MC-LR.
The influence of the pH was investigated in the pH
range 4–9 in experiments performed with similar initial
concentration of both chlorine and MC-LR (95 and
2.3mM, respectively). Fig. 3 shows the first-order plotfor this group of experiments. Again, the results suit a
pseudo-first-order kinetic model. From these experi-
mental values of k0, kapp was determined at each pH,
being the obtained values represented in Fig. 4
(symbols). It can be observed that the degradation of
MC-LR is favored at low pHs. These results are
explained if HOCl is the major electrophilic species
and the reactivity of OCl� is negligible. In addition, the
increase in the reaction rate in acidic solutions (pHo4)suggests that the reaction is subject to acid catalysis.
This pH-dependence was described by considering the
speciation of chlorine (reaction 3, pKa3 ¼ 7.5 (Albert
and Serjeant, 1984)), the acid-catalyzed reaction of
HOCl with MC-LR (reaction 4), and the reactions of
MC-LR with HOCl (reaction 5) and OCl� (reaction 6).
HOCl3OCl� þHþ Ka3, (3)
MC-LRþHOClþHþ! products k4, (4)
MC-LRþHOCl! products k5, (5)
MC-LRþOCl� ! products k6. (6)
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0
2
4
6
8
10
12
14
16
18
0 50 100 150 200 250[Cl2]t0, �M
k', x
103
s-1
Fig. 2. Plot or the pseudo-first-order rate constant versus the
initial concentration of chlorine: calculation of the apparent
second-order rate constant at 20 1C and pH 7.2.
0 50 100 150 200-3.0
-2.7
-2.4
-2.1
-1.8
-1.5
-1.2
-0.9
-0.6
-0.3
0.0
ln([
MC
-LR
]/[M
C-L
R] 0
)
time, s
Fig. 3. Influence of the pH on the pseudo-first-order rate
constant at 20 1C (pH ¼ 4.8 (’); pH ¼ 5.6 (&); pH ¼ 6.3 (K);
pH ¼ 7.2 (J); pH ¼ 8.0 (m); pH ¼ 8.8 (B)). Initial concentra-
tions of MC-LR and chlorine were 2.3 and 95mM, respectively.The solid line is a linear least-squares regression of the data.
1
10
100
1000
4 6 8 10pH
kapp
, M-1
s-1
5 7 9
Fig. 4. pH dependence of the apparent rate constants for the
reaction of chlorine with MC-LR. The line is calculated
according to Eq. (9).
J.L. Acero et al. / Water Research 39 (2005) 1628–16381632
From this reaction set, the degradation of MC-LR
can be written as follows:
�d½MC-LR�
dt¼ ðk4½HOCl�½H
þ� þ k5½HOCl�
þ k6½OCl��Þ½MC-LR�. ð7Þ
Taking into account the speciation of chlorine
(reaction 3), Eq. (7) is transformed into Eq. (8):
�d½MC-LR�
dt¼
ðk4½Hþ� þ k5Þ
½Hþ�
½Hþ� þ Ka3
� �þ k6
Ka3
½Hþ� þ Ka3
� �� �
� ½MC-LR�½Cl2�t, ð8Þ
where [Cl2]t ¼ [HOCl]+[OCl�]. Finally, by comparing
Eqs. (1) and (8), the apparent second-order rate constant
at each pH, kapp, is given by Eq. (9)
kapp ¼k4½H
þ�2 þ k5½Hþ� þ k6Ka3
ðKa3 þ ½Hþ�Þ. (9)
The individual rate constants were determined by a
non-linear least-squares regression of the experimental
pH profile of kapp (Eq. (9)). The obtained values were
2.07� 107, 116 and 6.78M�1 s�1 for k4, k5 and k6,
respectively. Fig. 4 depicts the experimental and
theoretical pH profiles of the apparent second-order
rate constants for the reaction of MC-LR with chlorine.
Reasonable good agreement between the experimental
values (symbols) and the calculated values from Eq. (9)
(lines) confirms the proposed reaction set. In general,
kapp decreases continuously in the pH range from 4 to 9.
These results indicate that the main reaction pathway at
circumneutral pH range is the reaction between MC-LR
and HOCl (reaction 5). In effect, the acid catalyzed
reaction becomes important at pHo5 and the rateconstant for reaction 6 is 20 times smaller than the
corresponding rate constant for reaction 5. Conse-
quently, acid catalyzed reaction of MC-LR with HOCl
(reaction 4) as well as the reaction of MC-LR with OCl�
(reaction 6) can be considered as negligible during
chlorination of natural waters (pH region 6–8).
Based on earlier chlorination studies of phenols (Lee,
1967; Rebenne et al., 1996), the above mechanism would
not be valid for acidic pH if high concentrations of
chloride ions were present. Under these conditions, Cl2can be formed from HOCl and Cl�. Since Cl2 is more
electrophilic than HOCl, its presence could accelerate
MC-LR degradation (Rebenne et al., 1996). The low
Cl� concentration in our commercial hypochlorite
reagent did not apparently affect the rate of MC-LR
ARTICLE IN PRESS
0.0 -0.5 -1.0 -1.5 -2.0
-2.0
-1.5
-1.0
-0.5
0.0
ln([
MC
-RR
]/[M
C-R
R] 0
)
ln([MC-LR]/[MC-LR]0)
Fig. 5. Competition kinetics plot for the calculation of the
apparent rate constant for the reaction of chlorine with MC-
RR at pH ¼ 6 (’), pH ¼ 7 (&) and pH ¼ 8 (K). Experi-
mental conditions given in Table 2.
J.L. Acero et al. / Water Research 39 (2005) 1628–1638 1633
degradation since the reaction rates were first-order with
respect to MC-LR even at low pH.
This pH-dependence of MC-LR chlorination confirms
the results reported by Nicholson et al. (1994). They found
that aqueous chlorine as the chlorinating agent was a
more efficient oxidant for the degradation of MC-LR in
comparison to calcium hypochlorite or sodium hypo-
chlorite, which have a basic character. These results were
attributed to the effect of pH, being toxin degradation
significantly reduced above pH 8. Tsuji et al. (1997) also
reported that a free chlorine dose of 0.7–2.8mg l�1 was
enough to remove MC-LR from distilled water if the pH
was kept below 7.2. Therefore the pH of the natural water
must be adjusted to below 7.5–8 during the chlorination
process to ensure adequate toxin removal.
The temperature effect on the apparent second-order
rate constant was investigated between the range
10–30 1C in experiments performed with similar con-
centration of MC-LR (2.3 mM) and chlorine (100mM) atpH 7.2. The obtained values for the apparent second-
order rate constants were 53.4, 60.2, 72.0 and
92.6M�1 s�1 at 10, 15, 20 and 30 1C, respectively. The
activation energy was calculated to be 20.1 kJmol�1.
Although the reaction was reaction-controlled according
to the values of the rate constants, the activation energy
was slightly above the value of 20 kJmol�1 proposed by
Pilling and Seakings (1995) for diffusion-controlled
reactions. This activation energy is lower than
58.5 kJmol�1 proposed by Lee (1967) for phenol
chlorination. Therefore, temperature does not exert an
important effect of MC-LR chlorination.
3.2. Kinetics of the reaction of chlorine with MC-RR and
MC-YR
The rate constants for the reactions of chlorine with
MC-RR and MC-YR were determined by competition
kinetics. This method has been successfully apply for the
calculation of rate constants for the reactions of organic
compounds with several oxidants such as bromine
(Acero et al., 2005), ozone (Gurol and Nekouinaini,
1984), UV radiation (Zamy et al., 2004) and OH radicals
(Haag and Yao, 1992). This procedure consists of the
simultaneous degradation of mixtures of two organic
compounds with similar chemical structure. One organic
compound acts as the reference compound, whose rate
constant with the oxidant is known; and the second one
being the target compound, whose rate constant for the
reaction with the oxidant must be determined. In the
present study, the target compound was MC-RR or
MC-YR, and the reference compound was MC-LR. The
application of this competition kinetic methods leads to
the following equation:
lnMC-RR½ �
MC-RR½ �0
� �¼
kapp MC-RR
kapp MC-LRln
MC-LR½ �
MC-LR½ �0
� �, (10)
where kapp MC�RR and kapp MC�LR represent the
apparent second-order rate constants for the reactions
of chlorine with the target compound (MC-RR or MC-
YR) and with the reference compound (MC-LR).
According to this Eq. (10), a plot of ln([MC-RR]/[MC-
RR]0) against ln([MC-LR]/[MC-LR]0) should be a
straight line whose slope is the ratio of apparent rate
constants. As kapp MC�LR is known, kapp MC�RR can
then be determined. Fig. 5 shows this plot for experi-
ments performed with MC-RR at different pH and
20 1C. As can be observed, points lie satisfactorily
around straight lines, confirming the fulfillment of
Eq. (10). By least squares regression analysis, the slopes
kapp MC�RR/kapp MC�LR were determined, these values
being detailed in Table 2. The apparent second-order
rate constant for the chlorination of MC-RR (Table 2)
was determined from the values of kapp MC�LR at the
investigated pHs. From these results, it can be deduced
that the reactivity of MC-LR and MC-RR at circum-
neutral pH are very similar.
Since pure MC-YR was not available, a similar
experiment was performed with microcystin extract
from Sujelow reservoir, in which MC-LR, -RR and
-YR were present. The initial toxin concentration ratio
in this extract was determined to be 1:0.9:1.2 measured
as MC-LR:MC-RR:MC-YR. This experiment was
performed at pH 7 and 20 1C with and initial chlorine
concentration of 100 mM and the initial toxin concentra-
tions given in Table 2. The plot of Eq. (10) lead a value
of the ratio kapp MC�YR/kapp MC�LR of 1.08 with
r240.99. Therefore, a value of 98.8M�1 s�1 is proposed
for the apparent second-order rate constant for the
reaction of chlorine with MC-YR at pH 7 and 20 1C.
The values of the chlorination rate constants are very
similar for the investigated microcystins. According to
ARTICLE IN PRESS
Table 2
Competition kinetics experiments for the determination of the apparent second-order rate constants for MC-RR and MC-YR
chlorination at 20 1C ([chlorine]0 ¼ 100mM)
pH [MC-RR]0 (mM) [MC-LR]0 (mM) kapp MC�RR/kapp MC�LR kapp MC�LR (M�1 s�1) kapp MC�RR (M
�1 s�1)
6.1 2.35 2.35 1.02 127.8 130.3
7.0 2.30 2.30 0.99 91.5 90.6
8.0 2.35 2.35 1.02 33.1 33.8
7.0* 2.51 2.17 1.08 91.5 98.8
*Experiment with MC-YR as target compound.
J.L. Acero et al. / Water Research 39 (2005) 1628–16381634
these results, chlorine likely attacks the conjugated
double bond of the Adda residue of microcystins. Since
the Adda moiety is the same in the three investigated
toxins, they show similar susceptibility to chlorine
oxidation at this point. In general, as the stereochem-
istry of toxic microcystins variants is similar in this part
of the molecule, it could be expected that all microcystin
species, including MC-LA, would react similarly with
chlorine. Similar behavior was observed for the ozona-
tion of microcystins (Newcombe et al., 2003).
Moreover, chlorine attack at the Adda group is
supported by the reaction products formed. Tsuji et al.
(1997) reported that the main degradation product was
dihydroxymicrocystin formed through the chloronium
ion at the conjugated diene of Adda, followed by
spontaneous hydrolysis. Other products may be its
stereoisomers and/or regioisomers, which were not
noxious.
3.3. Kinetics of the reaction of monochloramine with
MC-LR
The rate constant for the reaction of monochloramine
with MC-LR was also determined. In addition to the
importance of monochloramine as a final disinfectant, it
can be formed during the chlorination of ammonia-
containing waters (Vikesland et al., 2001). Therefore,
when the overall degradation of microcystin in a
chlorination process is to be investigated, reactions with
chlorine and monochloramine must be considered. For
that, the rate constant for the reaction with monochlor-
amine needs to be determined.
Experiments were performed at pH 8 and 20 1C in
presence of excessive monochloramine to observe an
appreciable degradation of MC-LR within several
hours. The rate of MC-LR degradation fits well
second-order kinetics, first order in the concentrations
of MC-LR and monochloramine. Since the degradation
of MC-LR was very slow, appreciable errors in MC-LR
concentration could be made. Moreover, the reaction
lasted several hours, time enough for monochloramine
self-decomposition (Vikesland et al., 2001), which makes
the consideration of constant monochloramine concen-
tration unlikely. Therefore, only an upper limit of
1.2� 10�2M�1 s�1 could be estimated for the second-
order rate constant for the monochloramination of
MC-LR. This result is in agreement with the weaker
oxidizing capability of monochloramine when compared
to both hypochlorous acid and hypochlorite (Glaze,
1990). Accordingly, monochloramine has been found to
be ineffective at oxidising microcystin in natural waters
(Nicholson et al., 1994; Hart et al., 1998). With this
insignificant rate constant, chloramine is not a suitable
oxidant for microcystin degradation from drinking
waters and plays an insignificant role in the degradation
of this group of toxins in the distribution system.
3.4. Significance of the reaction between chlorine and
microcystins in drinking water treatment processes
The efficiency of microcystin degradation during a
chlorination process depends on chlorine stability in the
water being treated, and on the reactivity of chlorine
towards the toxin. Chlorine stability varies with the
natural water quality parameters such as NOM,
ammonia and bromide concentrations and also with
the treatment conditions (pH and temperature) (Gallard
and von Gunten, 2002b). Therefore, to predict micro-
cystin degradation, in addition to the knowledge of
microcystin chlorination rate constants, chlorine decay
kinetics has to be characterized in each particular source
of water, for both oxidation pretreatment (high chlorine
demand) or final disinfection (reduced chlorine de-
mand). The chlorination treatment is deemed efficient
if microcystin concentration is reduced to below the
guideline value of 1mg l�1 (WHO, 1998).Based on the apparent second-order rate constants at
several pHs, the half-life times showed in Table 3 were
determined for MC-LR at a constant residual chlorine
concentration of 1mg l�1 (typical of the primary
disinfection) and 0.5mg l�1 (typical of the final disinfec-
tion). The degradation of MC-LR takes place in the
order of minutes at pH 6 and hours at pH48. As aconsequence, pH is a key variable in the chlorination
process and should be kept neutral or slightly acidic.
Therefore, when calcium or sodium hypochlorite are
ARTICLE IN PRESS
Table 3
Summary of MC-LR half-life values for chlorination at
different pHs. The initial concentration of chlorine (1 or
0.5mg l�1) was assumed to be constant
pH kapp (M�1 s�1) MC-LR half-life (min)
[chlorine]0¼ 1mg l�1
[chlorine]0¼ 0.5mg l�1
6 132.9 6.2 12.4
6.5 112.2 7.3 14.6
7 91.5 8.9 17.8
7.5 61.8 13.2 26.4
8 33.1 24.7 49.4
8.5 16.7 49.1 98.2
9 10.1 81.1 162.2
0
10
20
30
40
50
0 20 40 60
chlorine exposure (CT), mg l-1 min
[MC
-LR
], �g
l-1
Fig. 6. Degradation of MC-LR versus chlorine exposure.
Influence of the initial MC-LR concentration at pH ¼ 7 and
20 1C. [MC-LR]0 ¼ 50 mg l�1 (&); [MC-LR]0 ¼ 20mg l
�1 (J);
[MC-LR]0 ¼ 5 mg l�1 (B).
0
2
4
6
8
10
0 20 40 60 80
chlorine exposure (CT), mg l-1 min
[MC
-LR
], �g
l-1
Fig. 7. Degradation of MC-LR versus chlorine exposure at
pH ¼ 6 (&), pH ¼ 7 (J) and pH ¼ 8 (B). T ¼ 20 1C, [MC-
LR]0 ¼ 10 mg l�1.
J.L. Acero et al. / Water Research 39 (2005) 1628–1638 1635
used in high dose then the pH will need to be adjusted to
below 8 to guarantee the efficiency of the process.
Half-life values included in Table 3 were calculating
assuming a constant oxidant concentration during the
overall treatment process, which is not a realistic
situation. Instead, a parameter that includes the oxidant
decay within the reactor is the oxidant exposure.
Chlorine exposure (CT) is defined as the time-dependent
concentration of chlorine integrated over the reaction
time. To reach microcystin reduction in a chlorination
process, a certain chlorine exposure is needed. The
required CT value for achieving a certain degradation of
microcystin can be calculated from the following
equation in a batch process:
CT ¼� ln ½MC�=½MC�0
� �kapp
, (11)
where [MC] in the final concentration of microcystin
(guideline value of 1mg l�1), [MC]0 is the initial
concentration of microcystin and kapp is the apparent
second-order rate constant for the chlorination of
microcystin at a given temperature and pH. Note that
the widely used method for calculating CT10 has not
been applied since it leads to an underestimation of the
oxidant or disinfectant exposure in the reactor (Elovitz
et al., 2000).
The required CT value depends directly on the initial
concentration of microcystin in the natural water to be
treated or disinfected. Thus, Fig. 6 shows the degrada-
tion of MC-LR versus the CT value starting with
different initial concentrations of MC-LR at pH 7 and
20 1C. The initial microcystin concentrations were
selected within the range of MC presence in natural
waters. Spoof et al. (2003) reported microcystin con-
centrations up to 42.0 mg l�1 in Aland (SW Finland) in
2001. According to Fig. 6, in case of high toxin
concentration (410mg l�1) a chlorine exposure higherthan 30mg l�1min is required to match the guideline
value in the final water.
As discussed above, the pH exerts an important
influence on kapp, being the overall process favored at
low pHs. Fig. 7 shows the pH influence on the CT value
required for the degradation of 10 mg l�1 of MC-LR at20 1C. It can be observed that MC-LR is relatively easy
to eliminate at pH 6 or even 7. However, the required
chlorine exposure at pH 8 is above 80mg l�1min, which
can only be reached with high initial chlorine concen-
trations and elevated reaction times. In such conditions,
other problems such as trihalomethane and other
organic halogen formation have to be considered
(Gallard and von Gunten, 2002b).
Table 4 recompiles the CT values (calculated from
Eq. (11)) required for the degradation of MC-LR from
50 or 10 to 1mg l�1at several pHs (6–9) and temperatures(10–25 1C) in a batch or plug-flow reactor. Results
ARTICLE IN PRESS
Table 4
Chlorine exposure (CT values) for reducing microcystin
concentration to 1 mg l�1 at different pH and temperature in abatch or plug-flow reactor
pH [MC-LR]0(mg l�1)
CT-values (mg l�1min)
10 1C 15 1C 20 1C 25 1C
6 50 46.6 40.2 34.8 30.3
10 27.4 23.6 20.5 17.8
7 50 67.7 58.4 50.6 44.0
10 39.8 34.4 29.8 25.9
8 50 187.1 161.3 139.8 121.8
10 110.3 94.9 82.3 71.7
9 50 617.2 526.0 458.6 399.1
10 363.3 309.6 269.8 234.9
Table 5
Chlorine exposure (CT values) for reducing microcystin
concentration to 1mg l�1 at different pH and temperature in aCSTR
pH [MC-LR]0(mg l�1)
CT values (mg l�1min)
10 1C 15 1C 20 1C 25 1C
6 50 583.9 503.3 436.3 380.0
10 107.2 92.4 80.1 69.8
7 50 847.7 731.2 663.7 551.7
10 155.7 134.3 116.4 101.3
8 50 2347.5 2020.3 1751.8 1525.9
10 431.2 371.1 321.7 280.3
9 50 7731.1 6589.0 5740.9 4998.6
10 1420.0 1210.2 1054.4 918.1
J.L. Acero et al. / Water Research 39 (2005) 1628–16381636
indicate that low pH and low initial MC-LR concentra-
tion with elevated temperatures leads to the best CT
values. These CT values are comparable to those
required for achieving 99 percent (2 log) inactivation of
Giardia cyst and one to two order of magnitude higher
than those required for achieving 99.99 percent
(4 log) inactivation of viruses at pH 6 through 9 (CT
values for Giardia cyst (69mg l�1min) and viruses
(3–6mg l�1min) published by the USEPA (USEPA,
2003)). However, the pH influence is more pronounced
in MC-LR degradation than in bacteria or viruses
disinfection. As a consequence, degradation of micro-
cystins by chlorination at high pH must be carefully
controlled in order to meet the guideline value of 1mg l�1.The treatment of the reaction kinetics for the batch
experiments is mathematically equivalent to reaction
kinetics for an ideal plug-flow reactor (PFR), where
reaction time (batch reactor) and HRT (flow reactor) are
synonymous (Levenspiel, 1998). However, in drinking
water treatment practices, full-scale reactors have non
ideal flow behavior and can be considered hybrid of
PFR and completely stirred tank reactor (CSTR). The
PFR or CSTR represent the most and least efficient
reactor configurations, respectively (Levenspiel, 1998).
In the extreme case of ideal CSTR, the CT required to
eliminate microcystin can be calculated from Eq. (12)
CT ¼½MC�0=½MC�� �
� 1
kapp. (12)
Table 5 recollects the CT values required for the
degradation of MC-LR from 50 or 10 to 1mg l�1 atseveral pHs (6–9) and temperatures (10–25 1C) in a
CSTR. The CT values are almost one order of magnitude
higher than those corresponding with the PFR (Table 4).
These calculations are an example of the best and worst
case operating conditions for the degradation of MC-LR
and might serve for establishing safety chlorine dose in
oxidation or disinfection processes.
The chlorine dose needed to reach a certain CT value
in the reactor depends on the water quality parameters,
due to their influence on chlorine stability (Gallard and
von Gunten, 2002b). This is especially important in
surface water from rivers or eutrophic lakes, which
chemical/physical composition (e.g. temperature, pH,
DOM, ammonia, bromide) can change frequently
(Gallard and von Gunten, 2002b). In general, any
compound that reacts with chlorine (DOM, ammonia,
bromide) decreases the chlorine exposure, and therefore,
higher oxidant dose must be added to the water to reach
the required toxin elimination. Thus, higher chlorine
doses must be used in waters with high organic content
(Nicholson et al., 1994; Nicholson and Rositano, 1997).
In presence of ammonia, chloramines are formed, which
act as a chlorine sink and are not effective in microcystin
degradation. The presence of bromide is also important,
leading to the formation of hypobromous acid (HOBr)
which is a powerful oxidant, especially for phenolic
compounds (Acero et al., 2005). Kinetics of microcystin
bromination needs to be determined to assess the
importance of the bromination pathway in the overall
chlorination process. Therefore, it is necessary to
calibrate each water with respect to the rate of chlorine
depletion in order to determine the initial dose that leads
to the required CT value.
4. Conclusions
The reactions of MC-LR, -RR and -YR with chlorine
fit overall second-order kinetics over a wide pH range,
first-order with respect to both microcystin and chlorine.
It was found that an increase of pH has a negative effect
on the microcystin degradation rate. Apparent second-
ARTICLE IN PRESSJ.L. Acero et al. / Water Research 39 (2005) 1628–1638 1637
order rate constant for the chlorination of MC-LR at
20 1C decreased from 475M�1 s�1 at pH 4.8 to
9.8M�1 s�1 at pH 8.8. From these apparent second-
order rate constants, rate constants for the reactions of
MC-LR with aqueous chlorine species (hypochlorous
acid and hypochlorite) were evaluated. The results
indicate that the main reaction at circumneutral pH
range is the reaction between MC-LR and HOCl.
Similar reactivity with chlorine was found for MC-RR
and MC-YR, supporting that the main reactive site of
microcystin is the Adda moiety. These chlorination rate
constants can be utilized to determine the rate of
cyanotoxins degradation under typical water treatment
conditions. Half-life times ranged from minutes at pH 6
to hours at pH above 8 for a constant chlorine
concentration of 1.0–0.5mg l�1 (typical of oxidation
pre-treatment and final disinfection doses). Although
microcystin concentrations in our experiments were
much higher than expected for natural waters, the
chlorination rate constants determined in this study are
independent on the chlorine concentration and therefore
can also be applied in natural waters with much lower
microcystin concentrations.
The chlorine exposure (CT values) required to reduce
MC-LR concentration to 1mg l�1 in batch or PFR arecomparable to those required for achieving 99 percent
(2 log) inactivation of Giardia cyst, and one to two order
of magnitude higher than those required for achieving
99.99 percent (4 log) inactivation of viruses at pH 6
through 9. Chlorination is therefore a feasible option for
microcystin degradation during oxidation and disinfec-
tion processes if the pH of the water is kept below 8
during the treatment. However, monochloramine, which
can be formed during chlorination of ammonia-contain-
ing waters, is not capable of oxidizing microcystins.
These results are an important factor in process design
and the choice of the applied oxidants for the removal of
cyanobacterial toxins during drinking water treatment.
Acknowledgements
This work was supported by the European Commis-
sion under the Fifth Framework Programme, research
project TOXIC (EVK1-CT-2002-00107).
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