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Electrochemical Transformation of Trace Organic Contaminants in Latrine Wastewater Justin T. Jasper, Oliver S Shafaat, and Michael R Hoffmann Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b02912 • Publication Date (Web): 26 Aug 2016 Downloaded from http://pubs.acs.org on August 28, 2016
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Environmental Science & Technology
Electrochemical Transformation of Trace Organic Contaminants in Latrine Wastewater
2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41
Justin T. Jasper,1 Oliver S. Shafaat,2 Michael R. Hoffmann1,* 1
2
Environmental Science and Engineering California Institute of Technology Pasadena, California 91106
Division of Chemistry and Chemical Engineering California Institute of Technology Pasadena, California 91106
Submitted to Environmental Science and Technology June 10th 2016
*corresponding author: Contact information: e-mail:
[email protected]; phone: (626) 395-4391
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Abstract
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Solar-powered electrochemical systems have shown promise for onsite wastewater treatment in
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regions where basic infrastructure for conventional wastewater treatment is not available. To
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assess the applicability of these systems for trace organic contaminant treatment, test compound
46
electrolysis rate constants were measured in authentic latrine wastewater using mixed-metal
47
oxide anodes coupled with stainless steel cathodes. Complete removal of ranitidine and
48
cimetidine was achieved within 30 min of electrolysis at an applied potential of 3.5 V (0.7 A L-
49
1
50
80%) was achieved within 3 h of electrolysis. Oxidation of ranitidine, cimetidine, and
51
ciprofloxacin was primarily attributed to reaction with NH2Cl. Transformation of trimethoprim,
52
propranolol, and carbamazepine was attributed to direct electron transfer and to reactions with
53
surface-bound reactive chlorine species. Relative contributions of aqueous phase ·OH, ·Cl, ·Cl2-,
54
HOCl/OCl-, and Cl2 were determined to be negligible based on measured second-order reaction
55
rate constants, probe compound reaction rates, and experiments in buffered Cl- solutions.
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Electrical energy per order of removal (EEO) increased with increasing applied potentials and
57
current densities. Test compound removal was most efficient at elevated Cl- concentrations
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present when treated wastewater is recycled for use as flushing water (i.e., ~75 mM Cl-; EEO =
59
0.2-6.9 kWh log-1 m-3). Identified halogenated and oxygenated electrolysis products typically
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underwent further transformations to unidentifiable products within the 3 hr treatment cycle.
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Identifiable halogenated byproduct formation and accumulation was minimized during
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electrolysis of wastewater containing 75 mM Cl-.
). Removal of acetaminophen, ciprofloxacin, trimethoprim, propranolol, and carbamazepine (>
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Introduction Approximately 2.7 billion people worldwide lack access to water for conventional
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sanitation, wastewater treatment, and subsequent disposal.1 The lack of proper sanitation has
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been estimated to lead to millions of deaths per year due to water related disease.2 Onsite
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wastewater treatment provides an alternative strategy to protect human and environmental health
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in regions where it is not practical to build, maintain, or operate the infrastructure necessary for
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centralized wastewater treatment. Onsite wastewater electrolysis can provide rapid disinfection,3
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chemical oxygen demand removal,4 and nutrient removal.5,6 The wastewater can be treated to an
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extent that is suitable for recycling within an integrated toilet facility and waste treatment
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system.7
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The fate of trace organic contaminants during electrochemical onsite wastewater
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treatment, however, has not been evaluated to date. Trace organic contaminants that are
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recalcitrant during electrochemical treatment may accumulate as water is recycled within the
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system, potentially impacting aquatic ecosystems when the water is discharged.8–10 In addition,
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trace organic contaminants may be transformed to halogenated byproducts during electrolysis in
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the presence of chloride and bromide,11,12 which may be more toxic than their parent
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compounds.13
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Electrochemical trace organic contaminant transformation has been demonstrated under
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various conditions in synthetic and authentic wastewater.14–17 Non-active anodes (e.g., boron-
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doped diamond) generate weakly-absorbed hydroxyl radicals,18 which can rapidly mineralize
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trace organic contaminants in simple electrolytes,19–21 as well as in complex matrices such as
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municipal wastewater treatment plant effluents22,23 and reverse osmosis retentates.24,25 Despite
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their treatment efficacy, boron-doped diamond electrodes are often prohibitively expensive and
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produce toxic compounds such as bromate and perchlorate,26 making them impractical for many
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wastewater treatment applications.27
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Active electrodes, such as mixed-metal oxide anodes (e.g., IrO2, RuO2), are often less
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expensive than boron-doped diamond anodes. While active anodes may produce chlorate, they
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typically do not generate perchlorate.28 Active anodes are therefore an attractive option for
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wastewater treatment. Transformation of trace organic contaminants with active electrodes is
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significantly enhanced in the presence of chloride ion,29–31 which is oxidized to form reactive
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chlorine species such as Cl2 and HOCl/OCl- at the anode surface. In the presence of NH3, HOCl
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will be rapidly converted to less reactive chloramines. Chlorine radicals (i.e., ·Cl/·Cl2-) also have
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the potential to contribute to organic contaminant transformation.31 In addition to free ·Cl/·Cl2-
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(aq),
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species are especially important for onsite electrochemical treatment systems that recycle treated
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water for flushing, resulting in elevated Cl- concentrations in the range of 20-100 mM.
adsorbed chlorine radicals (·Clads) may transform organic chemicals.16 Reactive chlorine
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The purpose of this study was to evaluate the suitability of mixed-metal oxide anodes
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paired with base-metal cathodes for the removal of a suite of trace organic contaminants from
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latrine wastewater collected in an onsite wastewater treatment system. Test compounds were
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chosen to represent a range of reactivity with HOCl and chloramines. The importance of trace
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organic contaminant transformation pathways were evaluated, including direct electron transfer,
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reaction with ·OH, and reactions with free available chlorine (FAC; HOCl + OCl-), chloramines,
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Cl2, and chlorine radicals. Transformation products were identified in order to understand the
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effect of electrochemical operating conditions on product formation and to determine whether
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potentially toxic transformation products accumulated or underwent further transformation.
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Materials and Methods Materials. All reagents were purchased from Sigma Aldrich at the highest available purity. Solutions were prepared using ≥ 18 MΩ Milli-Q water from a Millipore system. Reaction Rate Constants for FAC, NH2Cl, and Cl2. Second-order rate constants for
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the reaction of test compounds with FAC (HOCl + OCl-) and NH2Cl were measured as described
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previously.32 Details are provided in the Supporting Information (SI) text.
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Second-order Cl2 reaction rate constants were measured in acidic solutions (0.1 M HCl;
115
pH ≈ 1.3) of HOCl. At this pH value, more than 90% of reactive chlorine was expected to be in
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the form of Cl2. Individual test compounds (20-200 nM) were added to solutions of Cl2
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(0.15-1.5 µM) with minimal headspace. Test compound and Cl2 concentrations were chosen to
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provide the most accurate measurement of reaction rate constants while ensuring that Cl2 was in
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excess by at least 7.5 fold. Samples were withdrawn within 1 min and added to borate buffer
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(100 mM) with Na2S2O3 quencher (45 mM). Rate constants were calculated based on first-order
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removal kinetics of the test compounds using measured steady-state total chlorine
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concentrations.
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Reaction Rate Constants for the Chlorine Radical Anion, ·Cl2-. Second-order rate
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constants for the reaction of test compounds with the dichlorine radical anion (·Cl2-) were
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measured by nanosecond transient absorption laser flash photolysis of solutions containing
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Na2S2O8 (25 mM), NaCl (100 mM), and test compounds (0-100 µM).33,34 Excitation at 266 nm
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(8 ns FWHM, 10 Hz repetition rate) produced ·SO4- that rapidly reacted with Cl- to produce ·Cl2-
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(~1 µM). The decay of ·Cl2- was monitored at 340 nm, log-normalized, and plotted versus test
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compound concentrations to determine second-order rate constants. Due to the low solubilities
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of the test compounds, reaction rate constants below 5 × 107 M-1 s-1 could not be measured.
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Further details are provided in the SI text.
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Second-order rate constants between ·Cl2- and latrine wastewater organic carbon were
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estimated as above, by measuring ·Cl2- decay rates in various latrine wastewater dilutions that
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were previously sparged to remove NH4+.
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Test Compound Electrolysis. Electrolysis was conducted with previously described
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mixed-metal oxide anodes (Ti/IrxTayO2/[Bi2O3]z[TiO2]1-z; Nanopac, South Korea)4 and stainless
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steel cathodes. Undivided electrode arrays were comprised of an anode (4.4 cm2) sandwiched
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between two cathodes of the same surface area (3 mm separation; Figure SI 1). The applied
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potential was held constant at 3.0, 3.5, or 4.0 V between the anode and cathodes using a
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potentiostat (Neware, China). To ensure that anodes were free of contamination, fresh anodes
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were used for each experiment and were preconditioned in Na2SO4 (15 mM) at 3.5 V for 30 min
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prior to use. Electrolysis solutions (70 mL) were stirred at 350 RPM in uncovered beakers.
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Samples (50 µL) were diluted by a factor of 10 with water to eliminate matrix effects and were
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quenched with Na2S2O3 (45 mM) to prevent further reactions after sampling. Free available
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chlorine (FAC; in the absence of NH4+) or total chlorine (in the presence of NH4+) concentration
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in the bulk solution was measured periodically during electrolysis experiments.
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Charge-normalized test compound removal rates (k′; C-1) were calculated from pseudo-
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first order removal rates (k; s-1) using the average experiment current (I; A) to correct for
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differences between electrolytes’ conductivities:
150 151 152
k′ =
k I
(1)
Chronoamperometric experiments were conducted to verify reactivity via direct electron transfer for select compounds. Details are provided in the SI text.
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Electrolysis Solutions. Authentic latrine wastewater was collected from a pilot-scale
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electrochemical toilet system located on the Caltech campus (Pasadena, CA). The on-site
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treatment system transferred public users’ waste to a settling tank (150 L), the supernatant of
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which was stored in a holding tank (1600 L; HRT ≈ 30 d) prior to batch electrochemical
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treatment. Treated water was recycled within the system as toilet flushing water. Wastewater
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for bench-scale electrolysis experiments was collected from the holding tank and filtered
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(0.45 µm) prior to use. Wastewater was amended with pharmaceuticals (1 µM) and the probe
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compound para-chlorobenzoic acid (pCBA; 100 µM). In several experiments, the latrine
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wastewater was modified before each test by adding NaCl (45 mM), tert-butanol (TBA; 500
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mM), or Na2S2O3 (30 mM).
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Buffered solutions (pH = 8.75; 20 mM Na2B4O7) contained NaCl (0-75 mM) and test
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compounds as described above. In select experiments, TBA (500 mM) was added to scavenge
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reactive intermediates.
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Analytical Methods. Total organic and inorganic carbon were measured using a TOC
167
analyzer (Aurora). NO3-, Cl-, Br-, PO43-, SO42-, and NH4+ were analyzed by ion chromatography
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(Dionex ICS 2000).35 Total and free chlorine were measured using commercially available kits
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(Hach) based on standard methods with N,N-diethyl-p-phenylenediamine (DPD).35 Total
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chlorine measurements in latrine wastewater may have included a small contribution (< 1%)
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from the relatively unstable NHBrCl and NH2Br, due to reaction between NH2Cl and Br-.36 Test
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compounds and transformation products were analyzed by an ultra-high performance liquid
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chromatography (Waters Acquity UPLC) system coupled to a UV-detector (Acquity PDA; for
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pCBA) and a time of flight mass spectrometer (Waters XEVO GS-2 TOF; for the test
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compounds). Details are provided in the SI text.
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Transformation Product Identification. Electrolysis reaction product identification
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was conducted under similar conditions as described above, except at elevated test compound
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concentrations (10 µM) in solutions containing: Na2SO4 (20 mM); NaCl (20 mM); or NaCl
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(20 mM) and (NH4)2SO4 (10 mM). Undiluted samples were collected over 10 min to 1 hr
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electrolysis experiments and quenched with Na2S2O3 (45 mM). Major transformation products
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were identified by comparing total ion chromatographs of the control electrolysis solutions (with
182
no added compounds) to actual test solutions. Additional transformation products were
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identified using MassLynx software (Waters), which identified peaks that were not noticeable in
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the total ion chromatograph. Tentative transformation product formulae and structures were
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determined based on accurate mass determinations, isotopic patterns, fragmentation patterns,
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comparison to the literature, and authentic standards, when available. Details are provided in the
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SI text.
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Results and Discussion Test Compound Rate Constants with Reactive Chlorine Species. Second-order
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reaction rate constants were measured to evaluate reactive chlorine species’ contributions to test
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compound electrolysis.
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FAC and NH2Cl. Test compound reaction rates with excess FAC and NH2Cl were
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measured at pH values similar to latrine wastewater (i.e., 8.7). At this pH value, OCl- was the
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dominant chlorine species during reaction with FAC. Test compounds followed pseudo first-
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order kinetics at rates similar to those available in the literature (Tables 1 and SI 1). The use of
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Na2S2O3 as a quenching agent in this study reduced the rapidly formed N-chlorinated
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ciprofloxacin intermediate during reaction between NH2Cl and FAC.37 Therefore, first-order
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ciprofloxacin transformation rate constants, which represent the decay of the N-chlorinated
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ciprofloxacin intermediate, were reported. As expected, these rate constants were not affected
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significantly by NH2Cl and FAC concentrations (Figure SI 2). Acetaminophen reaction rates
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were significantly faster with FAC than previously reported (5-10 times).32,38,39 The reaction rate
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constant of acetaminophen with NH2Cl at pH 8.7 was also almost 2 orders of magnitude higher
204
than previously measured.32 The second-order reaction rate constant between ranitidine and
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FAC was too fast to measure under the conditions employed, but was estimated to be greater
206
than 8000 M-1 s-1.
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Cl2. Except for ranitidine, the test compounds’ reaction rate constants with excess Cl2
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were much faster than with FAC (i.e., 1-5 orders of magnitude; Table 1) and too fast to measure
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accurately using the method employed. Reported rate constants with Cl2 are therefore only an
210
estimate. Nonetheless, measured Cl2 reaction rate constants were similar to those previously
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calculated and those predicted for acid-catalyzed reaction with FAC (Table SI 1).32,37,41–43
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·Cl2-. Second-order ·Cl2- reaction rate constants with the easily oxidized thioesters
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ranitidine and cimetidine were near-diffusion limited (Table 1).44 Trimethoprim and propranolol
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also reacted with ·Cl2- at near diffusion-limited rates (> 109 M-1 s-1), while the reaction rate with
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metoprolol was about an order of magnitude slower. Reaction rate constants between ·Cl2- and
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carbamazepine, acetaminophen, and ciprofloxacin were too slow to measure (i.e., < 5 × 107 M-1
217
s-1). This result was surprising for acetaminophen, which based on a linear free energy
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relationship for phenols and its weakly electron donating amide substituent, was predicted to
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react with ·Cl2- at a rate greater than 108 M-1 s-1.45
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The second-order ·Cl2- reaction rate constant with latrine wastewater organic matter was
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determined to be 1.9 ± 0.1 × 103 (mg L-1)-1 s-1, which was about an order of magnitude lower
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than reaction rate constants between ·OH and natural organic matter.46
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Electrolysis of Test Compounds in Latrine Wastewater. Latrine wastewater
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electrolysis at the relatively low current densities employed did not significantly alter wastewater
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pH, NH4+ concentration, or total organic carbon concentration (i.e., < 10% change; see Table 2
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for water characteristics). Conversely, electrolysis of trace organic test compounds in latrine
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wastewater collected from Caltech’s pilot-scale on-site toilet system resulted in greater than 80%
228
compound transformation within 3 h of treatment (charge density of 7.7 × 103 C L-1; Figure 1).
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Labile compounds were removed by greater than 90% within 30 min at 3.5 V (charge density of
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1.5 × 103 C L-1). Similarly rapid removal rates have previously been reported for electrochemical
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treatment of reverse osmosis water using mixed metal oxide anodes, with ranitidine and
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acetaminophen being removed significantly faster (> 90% removal by 4.3 × 102 C L-1) than
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trimethoprim, metoprolol, and carbamazepine (> 90% removal by 1.6 × 103 C L-1).47
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Effect of Voltage and [Cl-] on Test Compound Electrolysis. Increased applied potentials
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(3.0-4.0 V) and current densities (~0.4-1.3 A L-1) during electrolysis of latrine wastewater
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enhanced test compound electrolysis first-order rate constants (k; s-1). However, the increase
237
was less significant between 3.5 and 4.0 V for some compounds (Figure 2). Electrolysis of
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latrine wastewater amended with additional Cl- to simulate Cl- accumulation due to recycling
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flushing water (i.e., 30 mM as collected vs. 75 mM after amendment) enhanced test compound
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removal rates by 2-5 times (Figure 2). Increased Cl- concentrations have also been shown to
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enhance removal rates of chemical oxygen demand and benzoic acid under conditions similar to
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those employed in this study.4,28,48
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On a charge-normalized basis (k′; C-1), compound removal rates were similar over the
244
applied potential range tested with 30 mM Cl-. However, addition of Cl- to latrine wastewater
245
enhanced charge-normalized removal rates by 2-3 times (Figure 2). This was in contrast to
246
charge-normalized removal rates in buffered NaCl solutions, which were not significantly
247
affected by increasing Cl- concentrations, or in the case of ranitidine were reduced (Figure SI 3).
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Electrolysis Energy Efficiency. When normalized for energy consumption, test
249
compound removal was most efficient at lower applied potentials and current densities (Figure
250
3), suggesting that a moderate applied potential and current density (e.g., 3.5 V; ~0.7 A L-1) may
251
be a reasonable compromises for achieving both rapid and energy-efficient treatment of trace
252
organic compounds in latrine wastewater. Elevated Cl- concentrations (75 mM) resulted in a
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substantial reduction in the electrical energy per order removal (EEO) of test compounds (Figure
254
3). Measured values (0.2-6.9 kWh log-1 m-3) were comparable to EEO values for electrolysis of
255
personal care and household products in gray water (~1-13 kWh log-1 m-3)49 and trace organic
256
compound removal by cathodic H2O2 production/UV treatment in municipal wastewater
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(~3-7 kWh log-1 m-3).50 For rapidly transformed compounds, EEO values with 75 mM Cl- were
258
also similar to those for ozonation of trace organic compounds in municipal wastewater (~0.1-
259
0.3 kWh log-1 m-3).51
260
Test Compound Electrolysis Mechanisms. Trace organic compound transformation
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during electrolysis is typically ascribed to a combination of direct electron transfer and reaction
262
with reactive oxygen species (especially ·OH), FAC (especially HOCl), chloramines, chlorine
263
radicals (·Cl and ·Cl2-), and surface-bound species (·Clads).27,52 However, in most studies the
264
contributions of each of these mechanisms to contaminant removal is not evaluated. To provide
265
insight into the importance of the various transformation pathways during latrine wastewater
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electrolysis, reactive intermediates were measured and transformation rates were compared to
267
those measured in buffered Cl- solutions.
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·OH and ·CO3-. The selective ·OH probe pCBA was not removed during electrolysis in
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buffered solutions or in latrine wastewater (< 1% removal over 3 h; data not shown), implying
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that ·OH did not contribute to test compound transformation. This was in agreement with
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previous reports of low ·OH production for active mixed-metal anodes.18
272
The contribution of ·CO3- to test compound electrolysis could not be evaluated directly
273
due to the reactivity of ·CO3- probe compounds (e.g., N,N-dimethylaniline) with reactive chlorine
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species. However, reaction with ·CO3- is expected to be insignificant to the transformation of
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most of the compounds tested which exhibit relatively low ·CO3- reaction rate constants (i.e.,
276
carbamazepine, metoprolol, propranolol, trimethoprim; k·CO3- < 108 M-1 s-1).46 Transformation
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by ·CO3- may have been more important for compounds with higher reaction rate constants (e.g.,
278
acetaminophen; k·CO3- = 4 × 108 M-1 s-1).16
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Direct Electron Transfer. The contribution of compound oxidation via direct electron
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transfer (including reaction with adsorbed ·OH) during latrine wastewater electrolysis was
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evaluated by comparing charge-normalized transformation rates in buffered water to rates in
282
latrine wastewater. This was possible because free ·OH did not contribute significantly to test
283
compound removal (vide supra). Organic matter and ions present in latrine wastewater may
284
occupy active electrolysis sites on the surface. Thus, this method gave an estimate of the
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maximum contribution of direct electrolysis.
286
Test compounds transformation in buffered solutions followed first-order kinetics (Figure
287
SI 4). Removal rates were relatively slow (k′ < 8.3 × 10-4 C-1), except for ciprofloxacin,
288
ranitidine, and cimetidine (k′= (0.3-8.3) × 10-3 C-1; Figure 4). Oxidation via direct electron
289
transfer was confirmed for ranitidine and cimetidine chronoamperometrically (Figure SI 5).
290
Ciprofloxacin was not soluble enough for a similar effect to be observed. Compound adsorption
291
to the anode in the absence of applied current was only observed for the relatively hydrophobic
292
compounds propranolol and carbamazepine (log Kow > 2.5; Figure 4), and accounted for 27%
293
and 14% of their removal rates in buffered solutions, respectively.
294
Based on current-normalized rates in Cl--free buffered solutions, direct electron transfer
295
contributed significantly (~10-35%; Figure SI 6) to transformation of ciprofloxacin, cimetidine,
296
carbamazepine, and trimethoprim during latrine wastewater electrolysis. Direct electron transfer
297
contributed less than 10% to the removal of the other test compounds.
298
FAC. Electrolysis of Cl- produces Cl2 via the Volmer-Heyrovsky mechanism:53,54
299
MOx + Cl- → MOx(Cl·) + e-
(2)
300
MOx(Cl·) + Cl- → Cl2 + e-
(3)
301
Electrolysis of Cl- may also produce Cl2 via the Volmer-Tafel mechanism, with the second step
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303 304 305
2MOx(Cl·) → Cl2
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(4)
Rapid hydrolysis of Cl2 produces FAC,56 which may react with test compounds. As expected,29–31 test compound transformation rates in buffered solutions were
306
significantly enhanced by addition of Cl- (i.e., by 3-650 times versus solutions without Cl-;
307
Figure 4). Compounds exhibiting high reactivity with FAC (i.e., acetaminophen, cimetidine,
308
ranitidine; kFAC > 100 M-1 s-1) followed approximately second-order kinetics (Figure SI 7), and
309
their transformation could be attributed to reaction with accumulating FAC.29 Removal of
310
compounds with slower FAC reaction rate constants (propranolol, metoprolol, and
311
carbamazepine) followed first-order kinetics and reaction with FAC accounted for less than 35%
312
of their removal.
313
In latrine wastewater, FAC produced by Cl- electrolysis is expected to react rapidly with
314
NH3 (vide infra). Based on chlorine production rates estimated from initial chloramine formation
315
rates (1.7 × 10-6 M s-1) and the reaction rate constant between HOCl and NH3 (4.2 ×
316
106 M-1 s-1),57 steady state FAC concentrations were estimated to be less than 1 nM, which was
317
insignificant for test compound removal (< 0.2%). The observation that trimethoprim was the
318
slowest test compound to be removed during latrine wastewater electrolysis, despite its moderate
319
reactivity with FAC, supported this conclusion.
320
In wastewaters without NH3 (e.g., nitrified municipal wastewater effluent) FAC would be
321
expected to contribute significantly to transformation of test compounds with moderate to high
322
FAC reactivity. In wastewaters without NH3 that contain Br- (e.g., reverse osmosis concentrate),
323
reaction with HOBr also may contribute to trace organic compound transformation.16
324
NH2Cl. Chloramines produced during electrolysis of latrine wastewater accumulated
325
over the first 30 min of treatment, after which time their concentration remained approximately
326
constant due to chloramine reduction at the cathode (Figure SI 8). Chloramine production was 13 ACS Paragon Plus Environment
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consistent with the small NH4+ removal observed (i.e., ~1 mM; data not shown) at the relatively
328
low current density employed. Increased applied potentials and current densities during latrine
329
wastewater electrolysis, as well as increased Cl- concentrations, resulted in higher steady-state
330
chloramine concentrations (Figure SI 9). Reaction rate constants with NH2Cl were used to
331
assess test compound transformation with chloramines since NH2Cl is known to be the dominant
332
chloramine species at Cl2 to NH4+ molar ratios below 1 (in this study, Cl2:NH4+ ≈ 0.1-0.75 after 3
333
h treatment).58
334
Transformation of test compounds most reactive with NH2Cl (i.e., kNH2Cl > 0.9 M-1 s-1;
335
cimetidine, ranitidine, and acetaminophen) exhibited higher-order kinetics, suggesting their
336
removal was due to reaction with chloramines that accumulated during electrolysis (Figure 1).
337
Transformation of these test compounds also continued in unquenched samples at rates
338
comparable to those measured during electrolysis (data not shown), demonstrating that their
339
transformation was due to reaction with homogeneous reactants (e.g., chloramines). Based on
340
second-order NH2Cl reaction rate constants, removal of cimetidine and ranitidine could primarily
341
(>50%) be ascribed to reaction with NH2Cl (Figure SI 6). The combination of direct electron
342
transfer and reaction with NH2Cl accounted for greater than 70% of the removal of these
343
compounds. Acetaminophen also followed higher-order kinetics, even though only about 15%
344
of its removal could be attributed to reaction with NH2Cl. This suggested that acetaminophen
345
may have been removed by another reactive chlorine species that accumulated during
346
electrolysis such as NHCl2, which is known to be more reactive than NH2Cl.59 Ciprofloxacin
347
electrolysis rates at 3.5 V in wastewater, which reflected the decay of the rapidly formed N-
348
chlorinated ciprofloxacin intermediate (vide supra), were about 50% slower than predicted based
349
on measured first-order transformation rates with NH2Cl (Figure SI 6). This may have been due
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350
to reduction of the N-chlorinated intermediate at the stainless steel cathode, although further
351
experiments are necessary to verify this. Electrolysis rates of other test compounds followed
352
first-order kinetics, implying that their removal was not primarily due to reaction with
353
chloramines that accumulated during treatment. This was supported by their low second-order
354
reaction rate constants with NH2Cl, which accounted for less than 1% of test compound removal. Cl2. Although Cl2 generated at the anode is rapidly hydrolyzed (kCl2,H2O = 28.6 s-1),56
355 356
close to the anode transient Cl2 could theoretically contribute to test compound transformation.
357
However, compounds unreactive with NH2Cl were transformed at similar rates despite Cl2
358
reactivities spanning almost 3 orders of magnitude (e.g., compare metoprolol and trimethoprim;
359
Table 1). The possibility of mass transport limitations was eliminated by the increased
360
transformation rates observed with increased current densities under the same mixing conditions
361
(Figure 2). If rates were mass transport limited at current densities employed, then increasing the
362
current density would not substantially increase compound removal rates. Reaction with Cl2 was
363
therefore deemed not to be a significant transformation mechanism under the conditions
364
employed. ·Cl/·Cl2-. Organic compound electrolysis rates in Cl- solutions have previously been
365 366
correlated to reaction rates with ·Cl2-, implying that ·Cl2- contributed significantly to their
367
transformation.31 The chlorine radicals ·Cl and ·Cl2- can be formed via reaction of ·OH with Cl-
368
:60
369
∙ OH + Clି ↔∙ ClOH ି
370
∙ ClOH ି + H ା ↔∙ Cl + Hଶ O
371
∙ Cl + Clି ↔∙ Clି ଶ
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(5)
log K8 = 7.2
(6)
log K9 = 5.2
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·Cl has also been suggested to form by direct anodic Cl- oxidation.31 At Cl- concentrations
373
typical of latrine wastewater (i.e., > 30 mM Cl-), ·Cl2- is expected to have a steady-state
374
concentration almost 5000 times higher than that of ·Cl. However, second-order reaction rate
375
constants with ·Cl2- may be more than 3 orders of magnitude slower than with ·Cl, which are
376
typically similar to reaction rate constants with ·OH (i.e., > 109 M-1 s-1; Table 1).61
377
The contribution of dissolved ·Cl to test compound electrolysis in buffered Cl- solutions
378
was evaluated by comparing electrolysis rates with and without addition of TBA, which would
379
be expected to scavenge more than 99.9% of free ·Cl (k·Cl,TBA = 1.5 × 109).62 Based on the small
380
reductions in electrolysis rates (< 20%) observed in solutions amended with TBA (Figure 4), ·Cl
381
was determined not to contribute significantly to the electrolytic removal of most test
382
compounds.62 Quenching of dissolved ·Cl would also result in a reduction in ·Cl2- concentrations
383
by greater than 99.9% (eqn. 7), implying that ·Cl2- was not important for most test compounds’
384
removal in buffered Cl- solutions. Conversely, charge-normalized electrolysis rates for
385
propranolol, which exhibited a high reactivity with ·Cl2- but low reactivity with FAC (Table 1),
386
were reduced by more than 30% in Cl- solutions amended with TBA. Reaction with ·Cl2- may
387
therefore have been a significant transformation pathway for propranolol in buffered Cl-
388
solutions.
389
In latrine wastewater, quenching by organic matter is expected to result in even lower
390
steady state radical concentrations than in buffered Cl- solutions, suggesting that reaction with
391
·Cl2- did not contribute to test compound electrolysis. For example, based on the measured ·Cl2-
392
reaction rate constant with wastewater organic carbon (1.9 × 103 (mg L-1)-1 s-1), less than 1% of
393
·Cl2- is predicted to react with propranolol in the presence of 100 mg L-1 wastewater organic
394
carbon (see calculation in SI text). This conclusion was supported by a poor correlation between
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395
measured test compound ·Cl2- rate constants and observed test compound electrolysis rates in
396
latrine wastewater (Figure SI 10).
397
·Clads. Less than 20% of the electrolysis rates of carbamazepine, metoprolol, propranolol,
398
trimethoprim, and acetaminophen in latrine wastewater could be explained by the above
399
mechanisms (Figure SI 6). Acetaminophen removal was likely due to reaction with a
400
homogeneous species such as NHCl2 that accumulated throughout the reaction (vide infra).
401
However carbamazepine, metoprolol, propranolol, and trimethoprim exhibited first-order
402
kinetics (Figure 1) and their removal rates increased with increasing current densities (Figure 2).
403
This implied that they were transformed by a reactive species that rapidly reached a steady-state
404
concentration, and that compound electrolysis was kinetically-limited rather than mass transport-
405
limited at the current densities investigated.52 This conjecture was confirmed by the compounds’
406
stability in unquenched samples (data not shown), demonstrating that these test compounds were
407
transformed by reactive intermediates on or near the anode surface.
408
The most plausible transformation mechanism for carbamazepine, metoprolol,
409
propranolol, and trimethoprim that could not be quantified in this study was reaction with
410
surface-bound reactive chlorine species (i.e., ·Clads)48,63 formed during Cl- electrolysis (eqns. 2-
411
4). Surface-bound reactive chlorine species are expected to rapidly reach a steady-state surface
412
concentration that is enhanced with added Cl-,53,64 which agrees with the observed first-order
413
removal kinetics of carbamazepine, metoprolol, propranolol, and trimethoprim (Figures 1 and SI
414
4). It was not possible to evaluate the reactivity of ·Clads but it appeared to be inactive with TBA
415
(Figure 4), while it was efficiently quenched by addition of Na2S2O3 to electrolysis solutions.
416
Quenching by Na2S2O3 resulted in greater than an 80% reduction in test compound electrolysis
417
rates in latrine wastewater (Figure SI 11). Compound transformation on the anode surface may
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also explain the slightly higher electrolysis rate of propranolol compared with carbamazepine,
419
metoprolol, and trimethoprim. Propranolol was the most hydrophobic test compound and sorbed
420
most rapidly to the anode (i.e., ksorb ≈ 9 × 10-6 s-1 for propranolol vs. ksorb < 4 × 10-6 s-1 for other
421
compounds; Figure 4). Trace organic contaminant transformation via reaction with ·Clads has
422
also been suggested on Ti/IrO2 anodes in buffered solutions and municipal wastewater.16
423
Test Compound Transformation Products. Although electrolysis of latrine wastewater
424
provided efficient attenuation of test compounds, formation of potentially toxic transformation
425
products is cause for concern.27 This is particularly relevant as reactive halogen species (i.e.,
426
NH2Cl and ·Clads) were predominantly responsible for test compound transformation (vide
427
supra), and halogenation of aromatic pharmaceuticals may produce transformation products that
428
have a higher bioconcentration potential65 and are more toxic66,67 than their parent compounds.
429
For example, halogenated products formed during electrolysis of metoprolol in reverse osmosis
430
concentrate contributed significantly to an increase in toxicity of treated water.12 As a first step
431
in the evaluation of trace organic contaminant transformation products formed during
432
electrolysis of latrine wastewater, transformation products were identified and monitored during
433
wastewater electrolysis.
434
More than 50 test compound transformation products were identified with significant
435
responses during latrine wastewater electrolysis (Table SI 2). Identified transformation products
436
were typically hydroxylated and/or chlorinated and were similar to transformation products
437
formed during chlorination,32,37,39,42,43,68 chloramination,69 electrolysis,12 biological treatment,70–
438
73
439
the predominant propranolol transformation product (propranolol-Cl; m/z 294.1276 amu).
440
Further transformation product analysis is included in the SI text.
and oxidative treatment.74–76 Other products have apparently not been reported before, such as
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441
Transformation product yields in buffered Cl- solutions differed significantly from those
442
observed in latrine wastewater for test compounds reactive with FAC (Figure SI 12). For
443
example, hydroxylated and ketonated trimethoprim products were formed in significant
444
concentrations in buffered Cl- solution, but chlorinated products were favored in latrine
445
wastewater (Figure SI 12e). Similarly, many ranitidine transformation products observed during
446
latrine wastewater electrolysis were not observed during electrolysis of buffered Cl- solutions
447
(Figure SI 12g). This may have been due to rapid subsequent transformations in the presence of
448
FAC in Cl- solutions, as compared to slower reactions with chloramines formed in latrine
449
wastewater.
450
Test compound electrolysis products underwent additional transformation, in some cases
451
generating a succession of identifiable products (e.g., sequential decomposition of
452
ciprofloxacin’s piperazine group; Figure SI 13).37 At lower applied potentials and Cl-
453
concentrations, transformation products often accumulated during treatment, but under more
454
intense electrolysis conditions, transformation products were degraded further (Figure SI 13).
455
Electrolysis of latrine wastewater with 75 mM Cl- as compared to 30 mM Cl- actually resulted in
456
the formation and accumulation of fewer identified transformation products, including
457
halogenated products, over the 3 h treatment cycle (Figure 5). While removal of identified
458
halogenated transformation products during electrolysis is promising, without identification of
459
terminal trace organic contaminant electrolysis products it is not possible to evaluate the effect of
460
electrochemical treatment on trace organic contaminant toxicity. In latrine wastewater where
461
trace organic contaminants comprise only a small proportion of the organic matter (i.e., sub
462
µg L-1 trace organic compounds vs. ~100 mg L-1 bulk TOC), disinfection byproduct formation
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463
from bulk TOC is also likely to be of concern. Investigation of bulk TOC disinfection byproduct
464
formation during latrine wastewater electrolysis requires further study.
465 466
Acknowledgements
467
This research was supported by the Bill and Melinda Gates Foundation (BMGF RTTC Grant
468
OPP1111246) and a Resnick Postdoctoral Fellowship to JTJ. ·Cl2- reaction rate constants were
469
measured in the Beckman Institute Laser Resource Center at the California Institute of
470
Technology with funding provided by the Arnold and Mabel Beckman Foundation. We thank
471
James Barazesh and Cody Finke for useful discussions and critically reviewing the manuscript.
472 473
Supporting Information Available
474
Referenced Supporting Information, including additional materials and methods, discussion,
475
tables, and figures are provided free of charge via the Internet at http://pubs.acs.org.
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476
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Table 1. Test Compound Properties and Reaction Rate Constants. property compound
723 724 725 726
pKa
۽· ܓ۶(M-1s-1)
ܓ۴ۯ۱ (M-1s-1) a,b,c
ۼ ܓ۶ ۱( ܔM-1s-1) a,b
ܓ۱ܔ (M-1s-1) a
∙ ܓ۱ܔష (M-1s-1) a
metoprolol
9.6 (77)
8.4×109 (78)
1.9(±1.0)×10-2