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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

ACS Paragon Plus Environment


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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

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cimetidine was achieved within 30 min of electrolysis at an applied potential of 3.5 V (0.7 A L-

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1

50

80%) was achieved within 3 h of electrolysis. Oxidation of ranitidine, cimetidine, and

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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

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surface-bound reactive chlorine species. Relative contributions of aqueous phase ·OH, ·Cl, ·Cl2-,

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HOCl/OCl-, and Cl2 were determined to be negligible based on measured second-order reaction

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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

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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 =

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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 (>

1 ACS Paragon Plus Environment

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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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108 109 110 111


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;

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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

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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

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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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189 190


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

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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

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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

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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

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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%

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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

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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

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applied potential range tested with 30 mM Cl-. However, addition of Cl- to latrine wastewater

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enhanced charge-normalized removal rates by 2-3 times (Figure 2). This was in contrast to

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charge-normalized removal rates in buffered NaCl solutions, which were not significantly

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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

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compound removal was most efficient at lower applied potentials and current densities (Figure

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3), suggesting that a moderate applied potential and current density (e.g., 3.5 V; ~0.7 A L-1) may

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be a reasonable compromises for achieving both rapid and energy-efficient treatment of trace

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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

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3). Measured values (0.2-6.9 kWh log-1 m-3) were comparable to EEO values for electrolysis of

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personal care and household products in gray water (~1-13 kWh log-1 m-3)49 and trace organic

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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

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also similar to those for ozonation of trace organic compounds in municipal wastewater (~0.1-

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0.3 kWh log-1 m-3).51

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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

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with reactive oxygen species (especially ·OH), FAC (especially HOCl), chloramines, chlorine

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radicals (·Cl and ·Cl2-), and surface-bound species (·Clads).27,52 However, in most studies the

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contributions of each of these mechanisms to contaminant removal is not evaluated. To provide

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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

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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

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The contribution of ·CO3- to test compound electrolysis could not be evaluated directly

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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.,

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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

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latrine wastewater. This was possible because free ·OH did not contribute significantly to test

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compound removal (vide supra). Organic matter and ions present in latrine wastewater may

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occupy active electrolysis sites on the surface. Thus, this method gave an estimate of the

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maximum contribution of direct electrolysis.

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Test compounds transformation in buffered solutions followed first-order kinetics (Figure

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SI 4). Removal rates were relatively slow (k′ < 8.3 × 10-4 C-1), except for ciprofloxacin,

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ranitidine, and cimetidine (k′= (0.3-8.3) × 10-3 C-1; Figure 4). Oxidation via direct electron

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transfer was confirmed for ranitidine and cimetidine chronoamperometrically (Figure SI 5).

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Ciprofloxacin was not soluble enough for a similar effect to be observed. Compound adsorption

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to the anode in the absence of applied current was only observed for the relatively hydrophobic

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compounds propranolol and carbamazepine (log Kow > 2.5; Figure 4), and accounted for 27%

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and 14% of their removal rates in buffered solutions, respectively.

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Based on current-normalized rates in Cl--free buffered solutions, direct electron transfer

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contributed significantly (~10-35%; Figure SI 6) to transformation of ciprofloxacin, cimetidine,

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carbamazepine, and trimethoprim during latrine wastewater electrolysis. Direct electron transfer

297

contributed less than 10% to the removal of the other test compounds.

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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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in the reaction being:53,55 12 ACS Paragon Plus Environment


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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

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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,

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ranitidine; kFAC > 100 M-1 s-1) followed approximately second-order kinetics (Figure SI 7), and

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their transformation could be attributed to reaction with accumulating FAC.29 Removal of

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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%

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of their removal.

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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

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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.

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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

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FAC reactivity. In wastewaters without NH3 that contain Br- (e.g., reverse osmosis concentrate),

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reaction with HOBr also may contribute to trace organic compound transformation.16

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NH2Cl. Chloramines produced during electrolysis of latrine wastewater accumulated

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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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327

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

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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

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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

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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).

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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

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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

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electrolysis such as NHCl2, which is known to be more reactive than NH2Cl.59 Ciprofloxacin

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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


log K7 = -0.15

(5)

log K8 = 7.2

(6)

log K9 = 5.2

(7)

Page 17 of 35


372

·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



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


Page 18 of 35

Page 19 of 35


418

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



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


Page 20 of 35

Page 21 of 35


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.



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

Electrochemical Transformation of Trace Organic ... - ACS Publications

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