of 9
Toxic Byproduct Formation during Electrochemical Treatment of
Latrine Wastewater
Justin T. Jasper,
Yang Yang,
and Michael R. Ho
ff
mann
*
Environmental Science and Engineering, California Institute of Technology, Pasadena, California 91106, United States
*
S
Supporting Information
ABSTRACT:
Electrochemical systems are an attractive
option for onsite latrine wastewater treatment due to their
high e
ffi
ciency and small footprint. While concerns remain
over formation of toxic byproducts during treatment, rigorous
studies examining byproduct formation are lacking. Experi-
ments treating authentic latrine wastewater over variable
treatment times, current densities, chloride concentrations,
and anode materials were conducted to characterize by-
products and identify conditions that minimize their
formation. Production of inorganic byproducts (chlorate and
perchlorate) and indicator organic byproducts (haloacetic
acids and trihalomethanes) during electrolysis dramatically exceeded recommendations for drinking water after one treatment
cycle (
10
30 000 times), raising concerns for contamination of downstream water supplies. Stopping the reaction after
ammonium was removed (i.e., the chlorination breakpoint) was a promising method to minimize byproduct formation without
compromising disinfection and nutrient removal. Though treatment was accelerated at increased chloride concentrations and
current densities, byproduct concentrations remained similar near the breakpoint. On TiO
2
/IrO
2
anodes, haloacetic acids (up to
50
μ
M) and chlorate (up to
2
μ
M) were of most concern. Although boron-doped diamond anodes mineralized haloacetic
acids after formation, high production rates of chlorate and perchlorate (up to
4 and 25
μ
M) made them inferior to TiO
2
/IrO
2
anodes in terms of toxic byproduct formation. Organic byproduct formation was similar during chemical chlorination and
electrolysis of wastewater, suggesting that organic byproducts are formed by similar pathways in both cases (i.e., reactions with
chloramines and free chlorine).
INTRODUCTION
Onsite electrochemical systems show promise for providing
wastewater treatment to the billions of people lacking access to
adequate wastewater treatment,
1
and these systems are
currently being commercialized for application in both rural
communities (e.g., rural schools in South Africa) and urban
communities (e.g., apartment buildings in India). Electro-
chemical systems can be powered by solar energy and do not
require external water inputs, as treated water can be recycled
for
fl
ushing.
2
However, in addition to being recycled within the
system, once storage tanks are full, treated water is also
discharged to the environment due to system users
urine input.
Ensuring a high level of wastewater treatment is therefore
critical to protecting the receiving environment as well as
human health if discharged water reaches drinking water
sources or system users come in contact with recycled
fl
ushing
water.
Electrochemical treatment systems have been shown to
provide e
ff
ective treatment of latrine wastewater. Greater than
5-log inactivation of bacterial and viral indicator organisms is
achieved via production of reactive chlorine species from
chloride (15
20 mM) within 1 h at 4 V applied cell potential.
3
A combination of reactive chlorine species and direct oxidation
provide reduction of chemical oxygen demand (COD)
4
and
transformation of trace organic contaminants within 4 h (3.5
4.5 V applied cell potential)
5
7
with rates enhanced at elevated
chloride concentrations (up to 75 mM). Under similar
operating conditions, ammonium removal occurs via break-
point chlorination,
8
and phosphorus can be precipitated as
hydroxyapatite.
9
Unfortunately, strongly oxidizing conditions in the presence
of the high concentrations of chloride and organic matter
typical of latrine wastewater also result in formation of toxic
byproducts.
10
While wastewater disinfection is essential to
protecting human health, an ideal electrochemical system
should also be designed to minimize toxic byproduct
formation.
11
Chloride enhances electrochemical treatment via formation
of reactive chlorine species (e.g., hypochlorous acid, chlor-
amines, and chlorine radicals).
5
,
6
,
12
,
13
However, electrochemical
oxidation of chloride also produces the toxic byproducts
chlorate and, on
nonactive
anodes that preferentially form
hydroxyl radical (e.g., boron-doped diamond (BDD)),
Received:
March 23, 2017
Revised:
May 22, 2017
Accepted:
May 24, 2017
Published:
May 24, 2017
Article
pubs.acs.org/est
© 2017 American Chemical Society
7111
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
This is an open access article published under a Creative Commons Attribution (CC-BY)
License, which permits unrestricted use, distribution and reproduction in any medium,
provided the author and source are cited.
perchlorate.
10
,
14
,
15
For example, chlorate and perchlorate
concentrations 1000 to more than 100 000 times higher than
World Health Organization (WHO) and United States
Environmental Protection Agency (US EPA) health guidelines
were measured during electrochemical treatment of reverse
osmosis retentate,
16
,
17
urine,
18
and latrine wastewater
19
(see
Table SI 1
for health guidelines and
Table SI 2
for a summary
of previous electrochemical byproduct studies). Though less
toxic, nitrate may also be produced during electrochemical
treatment.
8
Electrochemical bromate production
20
is limited by
the low bromide concentrations typical of latrine wastewater.
5
A wide array of halogenated organic byproducts is also
produced by reaction of chlorine species with wastewater,
which contains organic compounds known to form disinfection
byproducts (i.e., carbohydrates, amino acids, and proteins).
21
Only a few indicator compounds such as trihalomethanes
(THMs) and haloacetic acids (HAAs) or the bulk parameter
adsorbable organic chlorine (AOCl) are commonly monitored,
however.
22
For example, electrolysis of latrine wastewater
3
and
reverse osmosis retentate
16
,
23
,
24
produced THMs and HAAs at
concentrations 10
50 times higher than those of drinking
water regulations (
Tables SI 1 and 2
).
25
Halogenated organic
byproducts have also been measured during treatment of
domestic wastewater,
26
,
27
latrine wastewater,
3
urine,
18
and
surface water.
28
While it is known that these toxic byproducts
are formed during electrochemical treatment, a comprehensive
study of the e
ff
ects of electrochemical operating conditions on
byproduct formation, with an aim of limiting byproduct
formation while ensuring adequate wastewater treatment, is
lacking.
The purpose of this study was for the
fi
rst time to rigorously
evaluate the formation of potentially hazardous byproducts
during electrochemical treatment of latrine wastewater. The
inorganic byproducts nitrate, chlorate, and perchlorate were
measured as well as the commonly observed organic
byproducts THMs (chloroform, bromodichloromethane, di-
chlorobromomethane, and bromoform) and chlorinated HAAs
(monochloroacetic acid (MCAA), dichloroacetic acid (DCAA),
and trichloroacetic acid (TCAA)), which were used as
indicators of halogenated organic byproduct formation. By-
product formation in authentic latrine wastewater was evaluated
under a range of treatment times, current densities, chloride
concentrations, and anode compositions. Operating parameters
were then identi
fi
ed that could be adjusted to minimize toxic
byproduct formation while maintaining e
ffi
cient wastewater
treatment.
MATERIALS AND METHODS
Materials.
All reagents were purchased from Sigma-Aldrich
at reagent-grade purity or higher. Solutions were prepared using
18 M
Ω
Milli-Q water from a Millipore system.
Latrine Wastewater.
Latrine wastewater was collected
from a previously described recycling electrochemical toilet
system located at Caltech (Pasadena, CA).
5
Wastewater from
an adjacent toilet was macerated and stored by the system,
treated electrochemically (3.5
4.0 V applied cell voltage; 4 h
batch treatment cycle), and then recycled for use as
fl
ushing
water. Latrine wastewater was collected from the untreated
water storage tank and
fi
ltered prior to use to enhance stability
between experiments (2.5
μ
m; see
Table 1
for water quality
parameters). Wastewater was amended with sodium chloride in
select experiments.
Wastewater Electrolysis.
Wastewater was treated electro-
chemically under conditions similar to those employed in the
Caltech electrochemical toilet system. Either mixed-metal oxide
anodes (TiO
2
/Ir
x
Ta
y
O
2
/Ti; referred to as TiO
2
/IrO
2
below for
simplicity; purchased from Nanopac, South Korea)
4
or BDD
anodes (purchased from NeoCoat, Switzerland) were coupled
to stainless steel counter electrodes. TiO
2
/IrO
2
anodes (14
cm
2
) were con
fi
gured as previously reported,
5
sandwiched
between two cathodes in 80 mL of wastewater. BDD anodes
(6.3 cm
2
) were paired with a single cathode in 25 mL of
wastewater as only one sid
e of the anode was active.
Electrochemical cells were undivided, and electrodes were
separated by 3 mm. Electrolysis current was held constant at
2.5
7.5 A L
1
(14
43 mA cm
2
, 3.6
4.4 V cell voltage for
TiO
2
/IrO
2
;15mAcm
2
, 6.5 V cell voltage for BDD) using a
potentiostat (Neware, China). Solutions were stirred (400
rpm) in uncovered beakers to simulate pilot-scale systems
which are left uncovered or actively vented to prevent
accumulation of hydrogen gas. The chlorination breakpoint
time (i.e., when ammonium removal was complete) was
identi
fi
ed by a peak in the measured voltage due to changes in
solution conductivity at the breakpoint (e.g.,
Figure SI 1
)as
well as by monitoring ammonium and total chlorine
concentrations.
Haloacetic Acid Electrolysis.
Electrolysis of individual
chlorinated HAAs (
1 mM) was evaluated in bu
ff
ered
solutions (30 mM sodium borate; pH 8.7) using TiO
2
/IrO
2
and BDD anodes. Electrolysis conditions were as described
above, except anodes were held at 7.5 V versus the cathode,
which was similar to the voltage measured during latrine
wastewater treatment with BDD anodes. In addition to HAA
concentrations, total organic carbon (TOC), chlorine, chloride,
chlorate, and perchlorate concentrations were measured to
determine if HAAs were mineralized during electrolysis. Para-
chlorobenzoic acid (pCBA) was added to select electrolysis
experiments (100
μ
M) to estimate hydroxyl radical steady state
concentrations.
Chlorination of Wastewater.
To compare byproduct
concentration pro
fi
les during chemical chlorination of latrine
wastewater to those observed during electrochemical treatment,
sodium hypochlorite (
5%) was added to stirred latrine
wastewater (20 mL) in uncovered beakers using a peristaltic
pump (0.47 mL h
1
). Sodium hypochlorite was added at a rate
similar to the initial rate of chlorine production during
electrochemical treatment (
16 mM h
1
).
Analytical Methods.
Total organic and inorganic carbon
concentrations were measured using a TOC analyzer (Aurora
Table 1. Typical Latrine Wastewater Properties
a
property
latrine wastewater
pH
8.6
±
0.2
conductivity (mS cm
1
)7
±
2
TOC (mg C L
1
)
b
154
±
13
TIC (mM)
c
52
±
2
COD (mg O
2
L
1
)
d
500
±
60
[Cl
] (mM)
e
33
100
[Br
](
μ
M)
5
±
1
[NH
4
+
] (mM)
34
±
2
a
Filtered latrine wastewater collected from Caltech onsite toilet.
b
Total organic carbon.
c
Total inorganic carbon ([HCO
3
]+
[CO
3
2
]).
d
Chemical oxygen demand.
e
[Cl
] was varied by NaCl
addition from 33 mM as collected to 100 mM.
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7112
1030; College Station, TX). Samples for ion analysis, including
ammonium, chloride, nitrate, chlorate, and perchlorate, were
diluted 25 times upon sampling and were quanti
fi
ed by ion
chromatography (Dionex ICS 2000; Sunnyvale, CA).
29
Total
chlorine and COD were measured within 5 min of sampling by
standard methods using commercially available kits (Hach;
Loveland, CO).
29
Organic disinfection byproducts were extracted immediately
upon sampling. THMs (1 mL sample aliquots) were extracted
using pentane (2 mL), and the organic phase was collected
using a glass transfer pipet for analysis following centrifugation
(5000 rpm, 5 min). HAA samples (1 mL) were amended with
Na
2
SO
4
(0.5 g), acidi
fi
ed (0.1 mL conc. H
2
SO
4
), and extracted
using methyl
tert
-butyl ether (2 mL).
30
HAAs were converted
to their corresponding methyl esters by addition of 10% H
2
SO
4
in methanol (1 mL) at 50
°
C (2 h). After conversion, samples
were cleaned with 10% Na
2
SO
4
in water (4 mL), and the
organic phase was collected for analysis.
HAAs and THMs were analyzed by gas chromatography
coupled to a mass spectrometer (GC/MS; HP 6890 GC/HP
5973 MS; Palo Alto, CA) in selected ion monitoring mode
(SIM) using previously reported methods that were modi
fi
ed
slightly.
30
,
31
Details are provided in the Analytical Methods
section of the
Supporting Information
text.
RESULTS AND DISCUSSION
Byproduct Formation during Wastewater Electrolysis.
With TiO
2
/IrO
2
anodes, electrolysis of latrine wastewater that
was amended with chloride (100 mM total Cl
) to simulate salt
accumulation during treated water recycling nearly completely
removed COD (
Table SI 3
).
4
Electrolysis also produced
chloramines, resulting in conversion of ammonium to nitrogen
gas (i.e., breakpoint chlorination;
Figure 1
).
8
Similar to
breakpoint chlorination via chlorine addition, about 5% of
ammonium was converted to nitrate (2 mM).
32
Prior to the breakpoint, hypochlorous acid reacted rapidly
with ammonia (4.4
×
10
6
M
1
s
1
) to form chloramines,
33
preventing the formation of chlorate on active anodes.
14
Following ammonium removal (
3.5 h), chlorate production
could be modeled as a series of reactions forming hypochlorous
acid and subsequently chlorate (
Figure SI 2
):
19
+→ +++
+→++
==
−−−+−
−−+−
−−−
−−
kk
2Cl H O OCl Cl 2H 2e
OCl 2H O ClO 4H 4e
0.35 M (Ah L ) ;
0.030 (Ah L )
k
k
2
23
1
111
2
11
1
2
Free chlorine concentrations reached a maximum of about 40
mM at 6 to 8 h, and the majority of chloride was converted to
chlorate by 12 h (75 mM). Oxidation of chlorate to perchlorate
was not detected on TiO
2
/IrO
2
anodes (i.e., <0.01 mM).
10
The
sum of chloride, chlorate, and total chlorine was always within
10% of the initial chloride concentration (
Figure 1
).
As expected, given the low bromide concentration (
Table 1
),
chloroform was the predominant measured THM, accounting
for greater than 95% of the total THMs measured (
Figure SI
3
). Chloroform concentrations increased until the breakpoint,
after which time concentrations decreased slowly (
Figure 1
).
Chloroform loss was attributed to volatilization,
18
as this loss
was also observed in the absence of applied current (data not
shown). Chloroform volatilization highlighted the need for
proper venting and possibly
fi
ltering the exhaust of electro-
chemical treatment systems.
18
In contrast to THMs, the nonvolatile chlorinated HAAs
accumulated throughout 12 h of electrolysis (
Figure 1
;
formation rates: MCAA: 0.5
±
0.1
μ
M (Ah L
1
)
1
; DCAA:
1.4
±
0.1
μ
M (Ah L
1
)
1
; TCAA: 0.6
±
0.1
μ
M (Ah L
1
)
1
).
HAAs were dominated by DCAA, which is typical during
chloramination of wastewater.
34
,
35
After the breakpoint when
free chlorine was present, TCAA formation rates increased
slightly (
25%), while DCAA and MCAA formation rates
remained constant or declined slightly. The increase in TCAA
formation rates may be explained by free chlorine
s preferential
production of TCAA.
36
Sustained accumulation of DCAA and
TCAA suggested that their organic precursors were not
signi
fi
cantly depleted during 12 h of electrolysis.
Maximum chlorate, chloroform, and HAA concentrations
were signi
fi
cantly higher (
10
50 times) than previously
measured during electrolysis of reverse osmosis retentate and
latrine wastewater (
Table SI 2
).
3
,
16
,
19
,
23
,
24
,
37
Higher production
of chlorinated byproducts compared to previous studies was
likely due to the higher initial concentrations of organic matter
(
4 times higher) and chloride (
2
25 times higher) in the
latrine wastewater used in this study (
Table SI 2
). Additionally,
the 12 h treatment time was excessive to achieve adequate
treatment. For example, chlorate concentrations prior to the
breakpoint (
2 h) were similar to those measured near the
breakpoint during electrolysis of reverse osmosis retentate
(
0.2 mM).
16
Minimizing the formation of toxic byproducts therefore
requires that electrolysis time be limited to the shortest time
necessary to achieve treatment goals. Chloroform was an
exception, as it was removed with further electrochemical
treatment. However, chloroform concentrations were generally
of less concern compared to drinking water regulations than
Figure 1.
Wastewater constituent concentrations and inorganic
byproduct concentrations (top) and organic byproduct concentrations
(bottom) during electrolysis (7.5 A L
1
; 3.8 V) of latrine wastewater
amended with Cl
([Cl
]
tot
= 100 mM) using TiO
2
/IrO
2
anodes.
Dotted lines indicate where the chlorination breakpoint was reached
(i.e., complete ammonium removal). Solid lines were added for clarity.
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7113
other byproducts (see below), and chloroform concentrations
would also be reduced post-treatment via volatilization in
treated water storage tanks.
Complete ammonium removal is a reasonable treatment goal
to protect the aquatic environment and prevent accumulation
of ammonium within the system, which can hamper treatment
e
ffi
ciency by scavenging reactive chlorine species.
5
Further,
disinfection and color remo
val are completed prior to
ammonium removal.
3
,
19
COD removal is typically more energy
e
ffi
cient before the breakpoint,
19
although achieving complete
COD removal may require pretreatment or longer treatment
times (
Table SI 3
). Halting treatment near the breakpoint will
also minimize residual free chlorine that may produce toxic
byproducts post-treatment (i.e., during storage of treated
water).
38
,
39
Electrolysis byproduct formation was thus evaluated
at the breakpoint under a variety of operating conditions to
optimize treatment while minimizing byproduct formation.
Current Density.
Ammonium removal rates increased
approximately proportionally with increasing current densities
with the breakpoint being reached at about 50 Ah L
1
at
current densities of both 5.0 and 7.5 A L
1
(the breakpoint was
not reached within 12.5 h at 2.5 A L
1
;
Figure SI 4
). Organic
and inorganic byproduct formation typically also increased in
proportion to current density (
Figures 2
and
SI 4
). Chlorate
formation, however, was reduced at lower current densities (i.e.,
65% lower concentrations after 60 Ah L
1
at 5.0 A L
1
than at
7.5 A L
1
), suggesting that chlorate formation is favored at
higher current densities and potentials.
19
TCAA formation was
also reduced at low current densities, possibly due to a smaller
near-anode concentration of free chlorine, which is thought to
form TCAA preferentially over DCAA.
36
Overall, while higher current densities may reduce treatment
times, similar concentrations of disinfection byproduct
concentrations will be formed at a given level of treatment
(e.g., degree of ammonium removal). The same conclusion was
previously reached for AOX production during treatment of
domestic wastewater.
27
Chloride Concentration.
As treated wastewater is recycled
in onsite treatment systems, chloride concentrations may
approach that of urine (i.e., 50 to >100 mM),
40
depending
on factors such as the extent to which additional water is added
by system users (e.g., hand washing or bidet water),
evaporation, and formation of halogenated byproducts. Higher
chloride concentrations may enhance electrolysis e
ffi
ciency due
to increased reactive chlorine species concentrations
5
but may
also be expected to increase chlorinated byproduct formation.
Increased chloride concentrations in latrine wastewater
resulted in higher steady-state total chlorine concentrations
(
0.5 mM with 30 mM Cl
;
0.8 mM with 65 mM Cl
;
1
mM with 100 mM Cl
) during electrolysis and thus more rapid
ammonium removal (i.e., the breakpoint was reached at 4.5 h
with 30 mM Cl
; 3.6 h with 65 mM Cl
; 2.8 h with 100 mM
Cl
;
Figure SI 5
).
Despite di
ff
erences in total chlorine concentrations, by-
product concentrations near the breakpoint were typically
within a factor of 2 (
Figures 3
and
SI 5
), as the breakpoint was
reached sooner with higher chloride concentrations. Nitrate
was an exception, as formation rates were similar at all chloride
concentrations throughout the electrolysis. Although organic
byproduct concentrations increased dramatically following the
breakpoint, after 6 h of treatment, they reached similar
concentrations at all chloride concentrations tested.
Figure 2.
Organic byproduct concentrations during electrolysis of
latrine wastewater at various current densities with TiO
2
/IrO
2
anodes.
Average cell voltages: 2.5 A L
1
: 3.6 V; 5.0 A L
1
: 4.0 V; 7.5 A L
1
: 4.4
V. Dotted lines indicate where the chlorination breakpoint was
reached (i.e., complete ammonium removal).
Figure 3.
Organic byproduct concentrations during electrolysis of
latrine wastewater at various chloride concentrations with TiO
2
/IrO
2
anodes. Average cell voltages: 30 mM Cl
: 4.4 V; 65 mM Cl
: 4.0 V;
100 mM Cl
: 3.9 V. Dotted lines indicate when the chlorination break
point was reached (i.e., complete ammonium removal).
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7114
Therefore, with the exception of nitrate, higher chloride
concentrations may enhance treatment e
ffi
ciency, but they
should not be expected to signi
fi
cantly a
ff
ect byproduct
concentrations near the breakpoint.
Anode Material.
In contrast to
active
oxygen-generation
anodes (e.g., TiO
2
/IrO
2
mixed-metal oxide),
nonactive
anodes (e.g., BDD) produce relatively high concentrations of
hydroxyl radical and low concentrations of hypochlorous acid.
19
This may result in mineralization of a greater proportion of
organic matter as opposed to accumulation as chlorinated
byproducts.
Electrolysis of latrine wastewater with BDD anodes produced
concentrations of toxic inorganic byproducts signi
fi
cantly
higher than those produced by treatment with TiO
2
/IrO
2
anodes. In contrast to TiO
2
/IrO
2
anodes, chlorate was formed
on BDD anodes throughout the electrolysis process, even in the
presence of ammonium (
Figure 4
). Chlorate was therefore
produced directly via oxidation of hypochlorous acid at the
anode,
41
whereas hypochlorous acid in the bulk solution
reacted rapidly with ammonia to form chloramines that are not
directly oxidized to form chlorate.
42
This was in contrast to
previous studies with Pt/Ti anodes in the absence of
ammonium that found direct oxidation of chloride to be only
a minor pathway for chlorate formation.
43
Chlorate was further oxidized to perchlorate, which
accounted for greater than 95% of the initial chloride
concentration (29 mM) after 6 h of electrolysis. Chloride,
chlorate, and perchlorate concentrations throughout electrol-
ysis could be
fi
t relatively well by a series of
fi
rst-order reactions
(
Figure SI 6
):
+→++
+→ ++
==
−−+−
−−+−
−−
−−
kk
Cl 3H O ClO 6H 6e
ClO H O ClO 2H 2e
0.17 (Ah L ) ;
0.087 (Ah L )
k
k
23
32
4
3
11
4
11
3
4
However, the model could not explain the lag in perchlorate
formation before 2 h. This lag may have been a result of initially
high chloride and organic matter concentrations, both of which
can inhibit electrochemical oxidation of chlorate.
44
Nonethe-
less, the sum of chlorine-containing species was within 10% of
the initial chloride concentrat
ion throughout electrolysis
(
Figure 4
).
Ammonium removal was limited to about 50% after 6 h of
electrolysis, as chloride oxidization to chlorate and perchlorate
competed with production of hypochlorous acid. Of the
ammonium that was removed, more than 60% was converted to
nitrate (
20 mM).
As on TiO
2
/IrO
2
, chloroform was the predominant
measured THM formed during electrolysis of latrine waste-
water on BDD anodes (>99% of THMs). Initial chloroform
formation rates on a charge density basis were similar on BDD
anodes and TiO
2
/IrO
2
anodes (
1
μ
M (Ah L
1
)
1
). However,
peak chloroform concentrations were reached much more
rapidly on BDD anodes (BDD at
6AhL
1
; TiO
2
/IrO
2
at
30 Ah L
1
). This may have been due to rapid mineralization
of organic precursors on BDD anodes (i.e., BDD: >90% TOC
removal; TiO
2
/IrO
2
:
30% TOC removal;
Table SI 3
). As on
TiO
2
/IrO
2
anodes, chloroform was volatilized following its
formation.
HAAs were initially formed at faster rates on a charge density
basis on BDD anodes as compared to TiO
2
/IrO
2
anodes (i.e.,
BDD:
1
4
μ
M (Ah L
1
)
1
; TiO
2
/IrO
2
:
0.5
1.5
μ
M (Ah
L
1
)
1
). This resulted in concentrations up to 2 times higher at
a similar level of treatment (
10 Ah L
1
). As on TiO
2
/IrO
2
anodes, DCAA dominated HAA production.
In contrast to treatment with TiO
2
/IrO
2
anodes, on BDD
anodes, chlorinated HAAs reached a peak concentration
between 1 and 2 h, after which time they were attenuated.
This implied that on BDD anodes, organic precursors were
removed, and HAAs were further oxidized. HAA attenuation
was pseudo
fi
rst-order (
R
2
= 0.94
0.99), and all HAAs were
removed at similar rates (MCAA: 12.7
±
0.4
×
10
2
(Ah
L
1
)
1
; DCAA: 8.2
±
1.0
×
10
2
(Ah L
1
)
1
; 8.4
±
2.5
×
10
2
(Ah L
1
)
1
;
Table SI 4
).
Electrolysis of HAAs in Borate Bu
ff
ered Solutions.
Electrolysis of individual chlorinated HAAs in borate bu
ff
ered
solutions con
fi
rmed that these compounds could be attenuated
on BDD anodes (
Figure 5
). As in latrine wastewater, removal
was pseudo
fi
rst-order (
R
2
> 0.99), and removal rates were 2
4
times higher than those observed in latrine wastewater
(MCAA: 26
±
1
×
10
2
(Ah L
1
)
1
; DCAA: 25
±
1
×
10
2
(Ah L
1
)
1
;35
±
1
×
10
2
(Ah L
1
)
1
;
Table SI 4
). This may
have been due to continued HAA formation in latrine
wastewater after attaining a peak concentration at 2 h.
Perchlorate formation and loss of TOC accounted for greater
than 95% of the initial HAA chlorine and carbon content,
suggesting that perchlorate and carbon dioxide were the
primary products of HAA electrolysis on BDD anodes.
Electrolysis rates were 50
125 times faster than expected for
reaction with hydroxyl radical (
k
·
OH,HAA
= (<6.0
9.2)
×
10
7
M
1
s
1
)
45
based on steady-state hydroxyl radical concen-
trations calculated using pCBA as a probe ([
·
OH]
ss
3
×
10
14
Figure 4.
Ion and inorganic byproduct (top) and organic byproduct
(bottom) concentrations during electrolysis (4 A L
1
; 6.5 V cell
voltage) of latrine wastewater with BDD anode. Lines added for
clarity.
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7115
M;
Figure SI 7
). Therefore, HAA electrolysis most likely
occurred via direct electron transfer at the BDD anode.
Conversely, on TiO
2
/IrO
2
anodes coupled to stainless steel
cathodes in borate bu
ff
ered solutions, TCAA and DCAA were
reduced to MCAA but not further transformed (
Figure 5
).
DCAA and TCAA could not be reformed from MCAA, as
released chloride was oxidized to chlorate. MCAA has also
previously been reported to be resistant to reduction on gold
and copper cathodes.
46
Chlorination of Wastewater.
With the exception of
MCAA, slow addition of hypochlorite solutions to latrine
wastewater (
16 mM h
1
) produced maximum halogenated
organic byproduct concentrations within a factor of 2 of
maximum concentrations measured during electrochemical
treatment with TiO
2
/IrO
2
anodes (
Figure 6
). When
normalized to percent ammonium removed (i.e., progress
toward the breakpoint), byproduct time pro
fi
les were also
similar (
Figure 6
).
The similarity in organic byproduct pro
fi
les suggested that
HAAs and THMs were formed by similar pathways during
electrolysis and chlorination, namely reaction with chloramines
and, after the breakpoint when ammonia was no longer present
to react with chlorine, free chlorine. As discussed above, the
predominance of DCAA prior to the breakpoint agreed with
previous studies showing that DCAA production is favored by
chloramines.
34
,
36
,
39
,
47
It is unclear why MCAA formation was
slower during chlorine addition than electrolysis of latrine
wastewater. One possibility is that MCAA production was
enhanced during electrolysis via cathodic reduction of DCAA
and TCAA (see above). Organic byproduct formation on BDD
anodes was also initially similar to byproduct formation during
chlorine addition, although byproducts were subsequently
removed on BDD anodes.
Chlorate concentration pro
fi
les during chlorine addition of
wastewater di
ff
ered dramatically from those during electrolysis.
During chlorine addition, chlorate accumulated linearly, as
chlorate was present in hypochlorite solutions as a decom-
position product (data not shown). Chlorate formation during
electrolysis was due to anodic oxidation of chloride and
hypochlorous acid following the breakpoint and was therefore
delayed.
Minimizing Electrochemical Byproducts
Health Im-
pacts.
As discussed above, complete ammonium removal
(breakpoint chlorination) is a reasonable goal when electro-
chemically treating latrine wastewater. Disinfection occurs well
before the breakpoint,
3
while formation of toxic byproducts is
generally minimized prior to the breakpoint. Stopping treat-
ment at the breakpoint may be achieved during batch operation
using an automated control system that monitors oxidation
reduction potential (ORP), which increases dramatically at the
breakpoint (
Figure SI 1
) and is measurable with robust and
inexpensive sensors.
To gain insight into the potential for byproducts of
electrochemical latrine wastewater treatment to contaminate
drinking water supplies, contaminant concentrations at the
breakpoint after one treatment cycle were compared to
drinking water guidelines (
Table SI 1
). Byproduct concen-
trations after treatment with TiO
2
/IrO
2
were typically 2 to 200
times above WHO drinking water guidelines. Nitrate was an
exception, as it was always below WHO guidelines (
Figure 7
).
Chlorate, MCAA, and DCAA posed the greatest risks to human
health, exceeding guidelines by more than 100 times. Trends
were similar when comparing byproduct concentrations to US
EPA drinking water limits and advisories (
Figure SI 8
),
Figure 5.
Haloacetic acid electrolysis using TiO
2
/IrO
2
(left) or BDD
(right) anodes. Solutions initially contained MCAA (top), DCAA
(middle), or TCAA (bottom) in borate bu
ff
ered solutions. Lines
added for clarity.
Figure 6.
Comparison of organic byproduct formation during addition
of hypochlorite to latrine wastewater and during electrochemical
treatment. Byproduct concentrations are plotted against the percent
NH
4
+
removed (i.e., progress toward breakpoint). Concentrations of
other chemical species are shown in
Figures 1
and
4
. Lines added for
clarity.
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7116
although byproduct to regulation ratios were higher (10
1000)
because EPA limits and advisories are generally more stringent
than WHO guidelines.
While treatment at low current densities or at high chloride
concentrations produced slightly lower byproduct concen-
trations, di
ff
erences were only within a factor of 2 to 3 at the
breakpoint. Changes in current densities and chloride
concentrations may therefore change treatment time and
energy e
ffi
ciency but do not substantially a
ff
ect byproduct
formation if treatment is stopped near the breakpoint.
BDD anodes oxidized chloride to chlorate and perchlorate
before complete ammonium removal was achieved. At the
point where chloride oxidation was nearly complete and
ammonium removal was maximized (i.e.,
4 h), organic
byproduct concentrations were signi
fi
cantly lower than those
during treatment with TiO
2
/IrO
2
anodes and were only about
20 times above WHO guidelines (
Figure 7
). However,
inorganic byproduct concentrations were much higher, with
perchlorate concentrations more than 10 000 times above
WHO guidelines.
Treatment with BDD anodes may instead be targeted toward
complete COD removal, which is more rapid than on TiO
2
/
IrO
2
anodes (
2h;
Table SI 3
), or removal of regulated
byproducts (
6 h). Even so, at these end points, chlorate and
perchlorate concentrations exceeded WHO guidelines by more
than 1000 times (
Figure SI 9
). Operating at lower current
densities may reduce perchlorate formation,
10
but this would
also necessitate an increase in treatment times and/or reactor
volume, further increasing the capital costs of using an already
expensive anode material. Mixed-metal oxide anodes such as
TiO
2
/IrO
2
are therefore preferable for their ability to limit
perchlorate and chlorate formation, despite their inability to
attenuate HAAs after formation. Bromate also may be produced
on BDD anodes,
20
although formation will be limited by the
low bromide concentrations typical of latrine wastewater (i.e.,
5
μ
M; maximum of
60 times WHO and EPA guidelines).
5
If electrochemically treated latrine wastewater is recycled as
fl
ushing water in an onsite system, discharged treated water will
undergo multiple treatment cycles depending on
fl
ushing water
volumes and other water inputs to the system (on average
about 11 cycles; see
Supporting Information
text for
calculation). With the exception of the volatile THMs,
byproduct concentrations in discharged water will therefore
be about 11 times higher than those after a single treatment
cycle, depending on variations between treatment cycles (e.g.,
chloride concentrations). Water discharged from onsite latrine
wastewater electrolysis systems will thus require additional
treatment before it can safely be used for human consumption.
The control of byproduct formation during electrochemical
treatment is complicated by the presence of both inorganic and
organic byproducts, although with additional research, certain
strategies may be e
ffi
cacious. Judicious siting of electrochemical
latrine wastewater treatment systems is a simple strategy which
may ensure discharged water is su
ffi
ciently diluted in drinking
water sources (more than 1000 times), thereby protecting
downstream consumers
health. A second strategy commonly
used during drinking water treatment
48
is to provide pretreat-
ment of latrine wastewater to remove organic byproduct
precursors.
49
Finally, use of novel electrode materials and
reactor designs may limit byproduct formation. For example,
activated carbon cathodes have been shown to capture and
reduce organic byproducts during treatment.
50
Alternatively,
latrine wastewater may be treated via reactive oxygen species
such as activated hydrogen peroxide produced at the cathode,
7
eliminating the formation of chlorinated byproducts. If
appropriately designed and operated, the dramatic decrease in
acute risk of disease provided by disinfecting latrine wastewater
will likely outweigh the long-term health implications of
chemical contamination of treated water.
51
ASSOCIATED CONTENT
*
S
Supporting Information
The Supporting Information is available free of charge on the
ACS Publications website
at DOI:
10.1021/acs.est.7b01002
.
Additional materials and methods, discussion, tables, and
fi
gures (
PDF
)
AUTHOR INFORMATION
Corresponding Author
*
Phone (626) 395-4391; fax: 626-395-2940; e-mail:
mrh@
caltech.edu
.
ORCID
Justin T. Jasper:
0000-0002-2461-5283
Yang Yang:
0000-0003-1650-4978
Notes
The authors declare no competing
fi
nancial interest.
ACKNOWLEDGMENTS
This research was supported by the Bill and Melinda Gates
Foundation (BMGF RTTC Grants OPP1111246 and
OPP1149755) and a Resnick Postdoctoral Fellowship to
J.T.J. We thank James Queen and Harry Collini for help with
sample analysis. We also thank James Barazesh, Eric Huang,
Figure 7.
Factors that byproduct concentrations near the chlorination
breakpoint ([byproduct]
BP
) exceeded WHO drinking water (DW)
guidelines with di
ff
erent anodes, current densities, and chloride
concentrations after one treatment cycle. For treatment with BDD
anodes, concentrations at complete chloride removal were used, as
complete ammonium removal was not achieved.
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7117
and Cody Finke for helpful discussion and critically reviewing
the manuscript.
REFERENCES
(1) Grant, S. B.; Saphores, J. D.; Feldman, D. L.; Hamilton, A. J.;
Fletcher, T. D.; Cook, P. L. M.; Stewardson, M.; Sanders, B. F.; Levin,
L. A.; Ambrose, R. F.; et al. Taking the
waste
out of
wastewater
for
human water security and ecosystem sustainability.
Science
2012
,
337
(6095), 681
686.
(2) Cho, K.; Kwon, D.; Hoffmann, M. R. Electrochemical treatment
of human waste coupled with molecular hydrogen production.
RSC
Adv.
2014
,
4
(9), 4596.
(3) Huang, X.; Qu, Y.; Cid, C. A.; Finke, C.; Hoffmann, M. R.; Lim,
K.; Jiang, S. C. Electrochemical disinfection of toilet wastewater using
wastewater electrolysis cell.
Water Res.
2016
,
92
, 164
172.
(4) Cho, K.; Qu, Y.; Kwon, D.; Zhang, H.; Cid, C. A.; Aryanfar, A.;
Hoffmann, M. R. Effects of anodic potential and chloride ion on
overall reactivity in electrochemical reactors designed for solar-
powered wastewater treatment.
Environ. Sci. Technol.
2014
,
48
(4),
2377
2384.
(5) Jasper, J. T.; Shafaat, O. S.; Hoffmann, M. R. Electrochemical
Transformation of Trace Organic Contaminants in Latrine Waste-
water.
Environ. Sci. Technol.
2016
,
50
(18), 10198
10208.
(6) Barazesh, J. M.; Prasse, C.; Sedlak, D. L. Electrochemical
Transformation of Trace Organic Contaminants in the Presence of
Halide and Carbonate Ions.
Environ. Sci. Technol.
2016
,
50
(18),
10143
10152.
(7) Barazesh, J. M.; Hennebel, T.; Jasper, J. T.; Sedlak, D. L. Modular
advanced oxidation process enabled by cathodic hydrogen peroxide
production.
Environ. Sci. Technol.
2015
,
49
(12), 7391
7399.
(8) Cho, K.; Hoffmann, M. R. Urea degradation by electrochemically
generated reactive chlorine species: Products and reaction pathways.
Environ. Sci. Technol.
2014
,
48
(19), 11504
11511.
(9) Cid, C.; Jasper, J. T.; Ho
ff
mann, M. Electrochemical phosphate
removal via precipitation of hydroxyapatite from human and domestic
wastewater.
Water Res.
,
2017
, in preparation.
(10) Radjenovic, J.; Sedlak, D. L. Challenges and Opportunities for
Electrochemical Processes as Next-Generation Technologies for the
Treatment of Contaminated Water.
Environ. Sci. Technol.
2015
,
49
(19), 11292
11302.
(11) Morris, J. C. Conference summary. In
Water chlorination
environmental impact and health e
ff
ects
; Jolley, R. L., Gorchev, H.,
Hamilton, D. H., Eds.; Ann Arbor Science: Ann Arbor, MI, 1978; Vol.
2
.
(12) Boudreau, J.; Bejan, D.; Bunce, N. J. Competition between
electrochemical advanced oxidation and electrochemical hypochlori-
nation of acetaminophen at boron-doped diamond and ruthenium
dioxide based anodes.
Can. J. Chem.
2010
,
88
(5), 418
425.
(13) Park, H.; Vecitis, C. D.; Hoffmann, M. R. Electrochemical Water
Splitting Coupled with Organic Compound Oxidation: The Role of
Active Chlorine Species.
J. Phys. Chem. C
2009
,
113
(18), 7935
7945.
(14) Tasaka, A.; Tojo, T. Anodic Oxidation Mechanism of
Hypochlorite Ion on Platinum Electrode in Alkaline Solution.
J.
Electrochem. Soc.
1985
,
132
(8), 1855
1859.
(15) Azizi, O.; Hubler, D.; Schrader, G.; Farrell, J.; Chaplin, B. P.
Mechanism of Perchlorate Formation on Boron-Doped Diamond Film
Anodes.
Environ. Sci. Technol.
2011
,
45
(24), 10582
10590.
(16) Bagastyo, A. Y.; Radjenovic, J.; Mu, Y.; Rozendal, R. A.;
Batstone, D. J.; Rabaey, K. Electrochemical oxidation of reverse
osmosis concentrate on mixed metal oxide (MMO) titanium coated
electrodes.
Water Res.
2011
,
45
(16), 4951
4959.
(17) Pe
́
rez, G.; Ferna
́
ndez-Alba, A. R.; Urtiaga, A. M.; Ortiz, I.
Electro-oxidation of reverse osmosis concentrates generated in tertiary
water treatment.
Water Res.
2010
,
44
(9), 2763
2772.
(18) Zo
̈
llig, H.; Remmele, A.; Fritzsche, C.; Morgenroth, E.; Udert,
K. M. Formation of Chlorination Byproducts and Their Emission
Pathways in Chlorine Mediated Electro-Oxidation of Urine on Active
and Nonactive Type Anodes.
Environ. Sci. Technol.
2015
,
49
(18),
11062
11069.
(19) Yang, Y.; Shin, J.; Jasper, J. T.; Hoffmann, M. R. Multilayer
Heterojunction Anodes for Saline Wastewater Treatment: Design
Strategies and Reactive Species Generation Mechanisms.
Environ. Sci.
Technol.
2016
,
50
(16), 8780
8787.
(20) Bergmann, M. E. H.; Iourtchouk, T.; Rollin, J. The occurrence
of bromate and perbromate on BDD anodes during electrolysis of
aqueous systems containing bromide: first systematic experimental
studies.
J. Appl. Electrochem.
2011
,
41
(9), 1109.
(21) Bond, T.; Goslan, E. H.; Parsons, S. A.; Jefferson, B. A critical
review of trihalomethane and haloacetic acid formation from natural
organic matter surrogates.
Environ. Technol. Rev.
2012
,
1
(1), 93
113.
(22) Hrudey, S. E. Chlorination disinfection by-products, public
health risk tradeoffs and me.
Water Res.
2009
,
43
(8), 2057
2092.
(23) Bagastyo, A. Y.; Batstone, D. J.; Kristiana, I.; Gernjak, W.; Joll,
C.; Radjenovic, J. Electrochemical oxidation of reverse osmosis
concentrate on boron-doped diamond anodes at circumneutral and
acidic pH.
Water Res.
2012
,
46
(18), 6104
6112.
(24) Bagastyo, A. Y.; Batstone, D. J.; Rabaey, K.; Radjenovic, J.
Electrochemical oxidation of electrodialysed reverse osmosis concen-
trate on Ti/Pt
IrO2, Ti/SnO2
Sb and boron-doped diamond
electrodes.
Water Res.
2013
,
47
(1), 242
250.
(25) US EPA.
Stage 1 Disinfectants and Disinfetion Byproducts Rule
,
EPA 816-F-01
014; U.S. Environmental Protection Agency: Wash-
ington, DC, 2001.
(26) Garcia-Segura, S.; Keller, J.; Brillas, E.; Radjenovic, J. Removal of
organic contaminants from secondary effluent by anodic oxidation
with a boron-doped diamond anode as tertiary treatment.
J. Hazard.
Mater.
2015
,
283
, 551
557.
(27) Schmalz, V.; Dittmar, T.; Haaken, D.; Worch, E. Electro-
chemical disinfection of biologically treated wastewater from small
treatment systems by using boron-doped diamond (BDD) electrodes
Contribution for direct reuse of domestic wastewater.
Water Res.
2009
,
43
(20), 5260
5266.
(28) Schaefer, C. E.; Andaya, C.; Urtiaga, A. Assessment of
disinfection and by-product formation during electrochemical treat-
ment of surface water using a Ti/IrO2 anode.
Chem. Eng. J.
2015
,
264
,
411
416.
(29) American Public Health Association.
Standard Methods for the
Examination of Water and Wastewater
, 19th ed.; American Public
Health Association, AWWA, Wat
er Environment Federation:
Washington, DC, 1995.
(30) Xie, Y. Analyzing Haloacetic Acids Using Gas Chromatography/
Mass Spectrometry.
Water Res.
2001
,
35
(6), 1599
1602.
(31) Weinberg, H. S.; Krasner, S. W.; Richardson, S. D.; Thruston, A.
D., Jr.
The Occurrence of Disinfection By-Products (DBPs) of Health
Concern in Drinking Water: Results of a Nationwide DBP Occurrence
Study
, EPA/600/R-02/068; U.S. EPA National Exposure Research
Laboratory: Athens, GA, 2002.
(32) Pressley, T. A.; Bishop, D. F.; Roan, S. G. Ammonia-nitrogen
removal by breakpoint chlorination.
Environ. Sci. Technol.
1972
,
6
(7),
622
628.
(33) Jafvert, C. T.; Valentine, R. L. Reaction scheme for the
chlorination of ammoniacal water.
Environ. Sci. Technol.
1992
,
26
(3),
577
586.
(34) Karan
fi
l, T.; Hong, Y.; Song, H. HAA Formation and Speciation
during Chloramination. In
Disinfection By-Products in Drinking Water,
ACS Symposium Series
; American Chemical Society, 2008; Vol.
995
,pp
124
140.
(35) Cowman, G. A.; Singer, P. C. Effect of Bromide Ion on
Haloacetic Acid Speciation Resulting from Chlorination and
Chloramination of Aquatic Humic Substances.
Environ. Sci. Technol.
1996
,
30
(1), 16
24.
(36) Diehl, A. C.; Speitel, G. E.; Symons, J. M.; Krasner, S. W.;
Hwang, C. J.; Barrett, S. E. DBP formation during chloramination.
Am.
Water Works Assoc. J. Denver
2000
,
92
(6), 76.
(37) Van Hege, K.; Verhaege, M.; Verstraete, W. Electro-oxidative
abatement of low-salinity reverse osmosis membrane concentrates.
Water Res.
2004
,
38
(6), 1550
1558.
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7118
(38) Hua, G.; Reckhow, D. A. Comparison of disinfection byproduct
formation from chlorine and alternative disinfectants.
Water Res.
2007
,
41
(8), 1667
1678.
(39) Goslan, E. H.; Krasner, S. W.; Bower, M.; Rocks, S. A.; Holmes,
P.; Levy, L. S.; Parsons, S. A. A comparison of disinfection by-products
found in chlorinated and chloraminated drinking waters in Scotland.
Water Res.
2009
,
43
(18), 4698
4706.
(40) Putnam, D. G.
Composition and Concentrative Properties of
Human Urine
, NASA CR-l802; NASA: Washington, DC, 1971.
(41) Czarnetzki, L. R.; Janssen, L. J. J. Formation of hypochlorite,
chlorate and oxygen during NaCl electrolysis from alkaline solutions at
an RuO2/TiO2 anode.
J. Appl. Electrochem.
1992
,
22
(4), 315
324.
(42) Kapa
ł
ka, A.; Joss, L.; Anglada, A
́
.; Comninellis, C.; Udert, K. M.
Direct and mediated electrochemical oxidation of ammonia on boron-
doped diamond electrode.
Electrochem. Commun.
2010
,
12
(12),
1714
1717.
(43) Jung, Y. J.; Baek, K. W.; Oh, B. S.; Kang, J.-W. An investigation
of the formation of chlorate and perchlorate during electrolysis using
Pt/Ti electrodes: The effects of pH and reactive oxygen species and
the results of kinetic studies.
Water Res.
2010
,
44
(18), 5345
5355.
(44) Bergmann, M. E. H.; Rollin, J.; Iourtchouk, T. The occurrence
of perchlorate during drinking water electrolysis using BDD anodes.
Electrochim. Acta
2009
,
54
(7), 2102
2107.
(45) Maruthamuthu, P.; Padmaja, S.; Huie, R. E. Rate constants for
some reactions of free radicals with haloacetates in aqueous solution.
Int. J. Chem. Kinet.
1995
,
27
(6), 605
612.
(46) Korshin, G. V.; Jensen, M. D. Electrochemical reduction of
haloacetic acids and exploration of their removal by electrochemical
treatment.
Electrochim. Acta
2001
,
47
(5), 747
751.
(47) Bougeard, C. M. M.; Goslan, E. H.; Jefferson, B.; Parsons, S. A.
Comparison of the disinfection by-product formation potential of
treated waters exposed to chlorine and monochloramine.
Water Res.
2010
,
44
(3), 729
740.
(48) Singer, P. C. Control of disinfection by-products in drinking
water.
J. Environ. Eng.
1994
,
120
(4), 727
744.
(49) Oller, I.; Malato, S.; Sa
́
nchez-Pe
́
rez, J. A. Combination of
Advanced Oxidation Processes and biological treatments for waste-
water decontamination
A review.
Sci. Total Environ.
2011
,
409
(20),
4141
4166.
(50) Li, Y.; Kemper, J. M.; Datuin, G.; Akey, A.; Mitch, W. A.; Luthy,
R. G. Reductive dehalogenation of disinfection byproducts by an
activated carbon-based electrode system.
Water Res.
2016
,
98
, 354
362.
(51) Regli, S.; Berger, P.; Macler, B.; Haas, C. Proposed decision tree
for management of risks in drinking-water: consideration for health
and socioeconomic factors. In
Safety of water disinfection: balancing
chemical and microbial risks
; Craun, G. F., Ed.; ILSI Press: Washington,
DC, 1993.
Environmental Science & Technology
Article
DOI:
10.1021/acs.est.7b01002
Environ. Sci. Technol.
2017, 51, 7111
7119
7119