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Effect of Ozone Treatment on NanoSized Silver Sulfide in Wastewater Effluent. Basilius Thalmann, Andreas Voegelin, Urs von Gunten, Renata Behra, Eberhard Morgenroth, and Ralf Kaegi Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.5b02194 • Publication Date (Web): 13 Aug 2015 Downloaded from http://pubs.acs.org on August 24, 2015

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Environmental Science & Technology

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Effect of Ozone Treatment on Nano-Sized Silver

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Sulfide in Wastewater Effluent.

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Basilius Thalmanna,b, Andreas Voegelina, Urs von Guntena,c,d, Renata Behraa,c, Eberhard

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Morgenrotha,b and Ralf Kaegia*

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a

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8600 Dübendorf

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b

ETH Zürich, Institute of Environmental Engineering, CH-8093 Zürich, Switzerland

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c

ETH Zürich, Institute of Biogeochemistry and Pollutant Dynamics, CH-8092 Zürich,

Eawag, Swiss Federal Institute of Aquatic Science and Technology, Überlandstrasse 133, CH-

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Switzerland

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d

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Fédérale de Lausanne (EPFL), CH-1015, Lausanne, Switzerland

School of Architecture, Civil and Environmental Engineering (ENAC), Ecole Polytechnique

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*

Corresponding author: [email protected]

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ACS Paragon Plus Environment

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ABSTRACT

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Silver nanoparticles used in consumer products are likely to be released into municipal

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wastewater. Transformation reactions, most importantly sulfidation lead to the formation of

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nanoscale silver sulfide (nano-Ag2S) particles. In wastewater treatment plants (WWTP),

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ozonation can enhance the effluent quality by eliminating organic micropollutants. The effect of

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ozonation on the fate of nano-Ag2S, however, is currently unknown. In this study we investigate

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the interaction of ozone with nano-Ag2S and evaluate the effect of ozonation on the short-term

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toxicity of WWTP effluent spiked with nano-Ag2S.

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The oxidation of nano-Ag2S by ozone resulted in a stoichiometric factor (number of moles of

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ozone required to oxidize one mole of sulfide to sulfate) of 2.91, which is comparable to the

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results obtained for the reaction of bisulfide (HS-) with ozone. The second order rate constant for

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the reaction of nano-Ag2S with ozone (k = 3.1×104 M-1 s-1) is comparable to the rate constant of

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other fast reacting micropollutants. Analysis of the ozonation products of nano-Ag2S by

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transmission electron microscopy (TEM) and X-ray absorption spectroscopy (XAS) revealed that

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ozonation dominantly led to the formation of silver chloride in WWTP effluent. After ozonation

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of the Ag2S-spiked effluent, the short-term toxicity for the green algae Chlamydomonas

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reinhardtii increased and reached EC50 values comparable to Ag+. This study thus reveals that

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ozone treatment of WWTP effluent results in the oxidation of Ag2S, and hence, an increase of

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the Ag toxicity in the effluent, which may become relevant at elevated Ag concentrations.

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KEYWORDS: Silver-Nanoparticles, Speciation, Ozone, Toxicity, Wastewater Treatment

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1.

Introduction

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Silver nanoparticles (Ag-NP) are incorporated into many consumer products,1 such as textiles

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and cosmetics2 because of their antimicrobial activity.3, 4 Several reports document that Ag-NP

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become detached from the bulk fabric during washing of the textiles and will thus be released to

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the sewer system.5-7 Results from mass flow analyses indicate that the largest fraction of Ag-NP

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will reach a wastewater treatment plant (WWTP).8 A high removal efficiency of ~95% has been

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observed for Ag-NP in WWTP. Consequently, ~5% of the Ag introduced as Ag-NP will reach

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the surface waters through the effluent of the WWTP.9

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In sewer systems and in WWTP, Ag-NP (including both metallic Ag- and AgCl-NP) are

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mainly transformed by sulfidation, resulting in nano-sized silver sulfide particles (nano-Ag2S).

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Only a minor fraction of the metallic Ag-NP, at least partially sulfidized, is expected to reach

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surface waters.10-14 Due to the ongoing reaction of Ag-NP with metal sulfides in urban surface

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waters, it is expected that the sulfidation of Ag-NP continues in oxygenated surface waters and

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will be completed within days to weeks.15 Ag2S shows a substantially reduced toxic response

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towards aquatic and terrestrial organisms compared to metallic Ag-NP,16-18 due to its very low

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solubility in aquatic systems. Ag2S is rather resistant towards oxidation and it remained stable for

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30 days in oxic waters19 and in sewage sludge during a composting period of 6 months.13

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Considerable loads of organic micropollutants pass WWTPs and negative impacts on the

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aquatic fauna and flora in the receiving waters are expected on a long-term.20 To mitigate the

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ecological impacts of micropollutants, an ozone treatment of the effluent water has been

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suggested to degrade the micropollutants.21, 22 In Switzerland, about 100 WWTP will, therefore,

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be upgraded with an additional treatment (ozone and powdered activated carbon (PAC)) stage

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and similar measures are currently discussed in Germany and other countries.


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In addition to the beneficial effect of degrading micropollutants, the treatment of effluent water

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with ozone may also affect the speciation of inorganic redox-sensitive compounds, including

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nano-Ag2S. The oxidative dissolution of metallic Ag-NP with ozone has been investigated in

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aqueous solutions.23, 24 The reaction of ozone with Ag2S, the most relevant form of Ag in treated

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wastewater, however, has not been addressed to date.

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The objectives of this study were, thus, to investigate the reaction of nano-Ag2S particles with

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ozone and to evaluate the effect of the ozone treatment on the Ag induced short-term toxicity of

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the WWTP effluent. We determined the stoichiometry for the reaction of ozone with nano-Ag2S

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and derived the second order rate constant for the reaction of nano-Ag2S with ozone by

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competition kinetics. The size, morphology and chemical speciation of the reaction products in

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effluent water were determined using transmission electron spectroscopy (TEM) and X-ray

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absorption spectroscopy (XAS) at the Ag K-edge. Furthermore, we assessed the toxicity of

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effluent water spiked with nano-Ag2S before and after ozonation by measuring the decrease of

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the photosynthetic yield of the freshwater algae Chlamydomonas reinhardtii exposed to the

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respective solutions.

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2.

Materials & Methods

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Starting Materials.

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Nano-Ag2S was synthesized by rapidly mixing 50 mL 3.7 mM silver nitrate solution (AgNO3,

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Merck) with 50 mL of a 1.9 mM sodium hydrogen sulfide solution (NaSH, Alfa Aesar) (both

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solutions prepared in doubly deionized water (DDI; Millipore, 18.2 MΩ cm) and with 10-5 M

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NaOH). This resulted in the immediate precipitation of nanoscale Ag2S. The suspension was

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sonicated for 20 min in a sonication bath (38 W L-1, Bioblock Scientific) to disperse the particles.

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For the synthesis no stabilizing agents (organics) were used due to possible reactions of the

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organics with ozone. The nano-Ag2S were collected for analysis by transmission electron

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microscopy (TEM), dynamic light scattering (DLS) and electrophoretic mobility measurements

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(zeta potential) as described in Thalmann, et al. (Table S1).15 The negative zeta potential (-39.8

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mV) of the nano-Ag2S particles prevented their agglomeration and the colloidal suspensions

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remained stable over several months.

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Ozone stock solutions (1.3-1.5 mM O3) were produced by bubbling ozone-containing O2 gas

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(produced with an ozone generator CMG 3-3, Innovatec) through deionized water (for all

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experiments doubly deionized (DDI) water was used, Millipore, 18.2 MΩ cm), which was

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cooled in an ice bath. The ozone concentration of the stock solution was determined in a UV-Vis

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spectrometer (ε = 3200 L mol-1 cm-1 at 260 nm, Cary 100, Varian).25

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Effluent water was obtained from a pilot WWTP (activated sludge process including primary

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clarification, nitrification, denitrification and secondary clarification) operated at Eawag.10 The

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WWTP effluent was characterized for dissolved organic carbon (DOC, Shimadzu TOC-L CSH),

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pH (SevenEasy, Mettler Toledo), chloride (Compact IC Flex 930, Metrosep A Supp 5 100/4.0,

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Metrohm), nitrate (Dr. Lange cuvettes LCK 339), ortho-phosphate (LCK 348) and chemical


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oxygen demand (COD, LCK 414). In addition, the concentration of copper (Cu) and zinc (Zn) in

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the effluent water was determined using inductively coupled plasma – optical emission

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spectroscopy (ICP-OES). Samples were filtered (GF/F filters, 0.7 µm, Whatman) before

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analysis, except for the pH measurements and for the elemental analysis.

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Stoichiometry of the nano-Ag2S-Ozone Reaction.

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The reaction of ozone with Ag2S results in the oxidation of sulfide to sulfate (SO42-), the

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production of O2 and the release of two Ag+ per SO42- (Eq. 1, up to 7.8 g Ag2SO4 L-1 is soluble in

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water).26 𝑘𝐴𝑔2𝑆

2𝐴𝑔+ + 𝑆𝑂42− + 𝑏𝑂2

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𝐴𝑔2 𝑆 + 𝑎𝑂3 →

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The stoichiometry of the reaction (b=(a×3-4)/2) can thus be determined by reacting known

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amounts of ozone with Ag2S and measuring the newly formed SO42-. Selected concentrations of

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ozone ranging from 0 to 100 µM were mixed with 12.5 µM nano-Ag2S. The pH of the solution

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was buffered with borate (pH 8.0, 10 mM). Tert-butanol (5 mM) was used as a radical scavenger

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to exclude unspecific side reactions through hydroxyl radicals.25 During the ozone addition, the

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solution was vigorously stirred. The amount of nano-Ag2S reacting with ozone was derived from

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the total sulfate concentration measured by ionic chromatography after completion of the

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reaction (IC; ICS3000, AS9HC, Dionex, eluent: 9 mM Na2CO3 / 2 mM NaOH, QL = 5 µg L-1).

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Samples for IC analysis were prepared by filtration through 10 kDa centrifugal membranes

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(Vivaspin 20, santorius stedim biotech, 10 min, 3100 g), followed by the removal of Ag+ from

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the filtrates with OnGuard II H-cartridges (Dionex). Dissolved ( 1.5 minimum particle

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diameter; ~60% of 105 analyzed particles) with dimensions of 57.5±15.2 × 19.3±3.9 nm and

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spherical Ag2S nanoparticles with a diameter of 25.1±9.3 nm (Figure S1). A more detailed

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summary of the particle characteristics is given in Table S1. The characteristics of the effluent

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water are given in Table S2.

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Stoichiometry of the Reaction of Nano-Ag2S with Ozone.

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To determine the stoichiometry of the reaction of Ag2S with ozone, the sulfate concentrations

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resulting from the oxidation of the sulfide were plotted against the applied ozone dose (Figure

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1A). The sulfate increased with increasing ozone dose and reached a plateau of ~12.5 µM sulfate

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at 35.6 µM ozone, corresponding to the maximum amount of sulfate that can be produced from

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12.5 µM of Ag2S. This indicates that at an ozone concentration of 35.6 µM, all Ag2S-sulfur has

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been oxidized to sulfate. Linear regression of the sulfate concentration against the ozone

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concentration (excluding the last measurement point where ozone was present in excess),

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resulted in a straight line with a slope of 0.343(5) (r2 = 1.0, standard error (1σ) in brackets). This

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slope corresponds to moles of SO42- molecules produced by the reaction of Ag2S with one mole

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ozone. Thus, 2.91 moles ozone were required to oxidize 1 mole S(-II) in Ag2S to S(+VI) (Eq. 1,

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a = 2.91, b = 2.37).


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Rate Constant for the Reaction of Nano-Ag2S with Ozone.

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The rate constant of the reaction of nano-Ag2S with ozone was determined from competition

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kinetics with buten-2-ol as competitor. Formaldehyde, the oxidation product of buten-2-ol, was

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measured. The rate constant of the reaction was derived according to Eq. 3 by linear regression

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(Figure 1B; for details see SI). The slope of the regression line (0.40(1)) corresponding to the

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ratio of the second order rate constants kBu/kAg2S was used to calculate kAg2S (3.1(1)×104 M-1 s-1,

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Eq. 2).

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Ozone Treatment of WWTP Effluent Spiked with Nano-Ag2S.

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During the ozonation of effluent, DOM oxidation will be in competition to the oxidation of

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nano-Ag2S. To study the influence of DOM on the extent of the oxidation of Ag2S, experiments

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were conducted in a simplified system with DDI water spiked with 2.8 µM Ag2S in the presence

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of humic acid as a proxy for DOM. Both the Agfree fraction and the AgNH3 fraction representing

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the sum of the “free” fraction and NH3-soluble Ag species, reached the concentration expected

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for total oxidation of the Ag2S already at the lowest experimental ozone dose (0.24 gO3 / gDOC

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corresponding to 20 µM O3, Figure 2A).

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To investigate the effects of anions (e.g. Cl- and PO43-) on the speciation of Ag after ozonation,

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experiments with humic acid, Cl- and PO43- were performed. The AgNH3 fraction showed an

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almost identical trend as observed in the absence of these anions (Figure 2B). However, the

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Agfree fraction was only elevated at the lowest specific ozone dose and then gradually decreased

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to below the detection limit at a specific ozone dose of 1 gO3 / gDOC.

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To assess the impact of an ozone treatment on the fate of Ag2S under WWTP-relevant

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conditions, nano-Ag2S (2.8 µM and 11.5 µM) was spiked to WWTP effluent and the speciation

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of Ag was investigated as a function of the specific ozone dose (Figure 2C and D). At lower


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nano-Ag2S concentrations, the AgNH3 fraction increased steadily with specific ozone doses and

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reached the concentration corresponding to the total oxidation of Ag2S at a specific ozone dose

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of 1 gO3 / gDOC. The Agfree fraction reached a maximum between 0.4 and 0.6 gO3 / gDOC (Figure

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2C). At higher nano-Ag2S concentrations the AgNH3 fraction only increased at O3 doses > 0.4 gO3

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/ gDOC and did not reach the expected Ag concentration based on a total oxidation of the spiked

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Ag2S. The Agfree fraction only slightly increased with specific ozone dose but always stayed

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below 1 µM (Figure 2D).

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To determine the Ag speciation in the ozone treated WWTP effluent bulk suspension in more

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detail and to better characterize the AgNH3 fraction, Ag K-edge XAS was performed. The

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EXAFS spectra of the ozone treated samples and of reference materials are given in Figure 3. Ag

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speciation was assessed by a LCF analysis using the spectra of Ag2S and AgCl as references.

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The resulting fractions using the EXAFS spectra for the linear-combination fitting (LCF)

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analysis are given in Figure 2D and comparable results were also obtained with the XANES

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spectra (Figure S3 and S4). The increase of the AgCl fraction with increasing ozone dose closely

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matched the increase of the AgNH3 fraction, supporting the assumption that NH3 allowed to

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selectively probe dissolved Ag and AgCl but not Ag2S (Figure 2D).

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TEM analysis of samples collected from the effluent water after ozonation revealed the

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formation of cubic particles with an equivalent spherical diameter of 50 nm – 200 nm (Figure 4).

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EDX analysis of selected cubes clearly indicated Ag and Cl and no detectable S (Figure 4 inset)

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confirming that the cubes represented AgCl particles that most probably formed after the release

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of Ag+ by Ag2S oxidation.

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Effect of the Ozonation on the Toxicity of Nano-Ag2S Spiked WWTP Effluent.


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Increasing concentrations of nano-Ag2S (up to 175 µM) spiked to WWTP effluent did not

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show any influence on the algal photosynthetic yield (n = 3). However, ozonation of effluent

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water spiked with increasing amounts of Ag2S reduced the photosynthetic yield (Figure 5). The

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dose-response curve was described with a logistic model (Table 1). The EC50 value increased

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from 0.256(15) µM after 1 h exposure time to 0.375(31) µM after 2 h exposure time (n = 3,

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Table 1). EC50 values obtained for AgNO3-spiked effluent water were comparable to the EC50

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values obtained for nano-Ag2S-spiked effluent after ozonation (Table 1, Figure 5).


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Discussion

Oxidative Dissolution of Nano-Ag2S by Ozonation.

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Ozone is a strong oxidant with a maximum of two accepted electrons for oxygen transfer

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reactions.22 Thus, for the oxidation of sulfide (-II) to sulfate (+VI), releasing eight electrons, four

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ozone molecules would be required. However, experimental observations of the oxidation of

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bisulfide (HS-) with ozone suggest that only 2.4 ozone molecules are needed to oxidize one

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sulfide to sulfate. Density functional theory (DFT) investigations of the reaction of sulfide with

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ozone identified two major reaction pathways. One pathway needs the postulated four ozone

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molecules, but the other requires only two ozone molecules as sulfinic peracid (HS(O)OO-) is

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formed as an intermediate and, therefore, more than two electrons can be transferred to one

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ozone molecule.37 We found that 2.91 mole of ozone are required to fully oxidize the sulfide

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from one mole of nano-Ag2S. Hence, slightly more ozone is needed to oxidize the Ag-bound

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sulfide compared to the HS-, but still significantly less than predicted based on the assumption of

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a two electron transfer per ozone molecule.

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Increasing ozone concentrations also resulted in increasing Ag+ concentration but only at

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ozone concentrations > ~15 µM (Figure S2). Furthermore, Ag+ concentration did not exceed 20

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µM (25 µM would have been expected for complete Ag2S oxidation) which most probably

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reflected losses of Ag+ to the walls of the reaction tubes.

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The oxidation of bisulfide with ozone (k = ~109 M-1 s-1; t1/2 = 3.5×10-5 s, assuming pseudo-

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first-order kinetics with [O3] = 20 µM)38 is much faster than the oxidation of nano-Ag2S (k =

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3.1×104 M-1 s-1, t1/2 = 1.1 s). The slight deviation of the measurements from linearity (Figure 1B)

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may be taken as an indication that the reaction rate was limited by mass transfer (diffusion). We

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conducted our experiments with nano-Ag2S that included individual particles close to 100 nm. If


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the reaction is mass-transfer limited, then smaller nano-Ag2S will be oxidized even faster during

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the ozone treatment of the wastewater effluent. For the oxidation of nano-Ag2S in effluent water

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the rate constant was comparable to bisphenol A (1.7×104 M-1 s-1, t1/2 = 2.0 s, pH = 7), an

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endocrine disruptor found in wastewater.39 Even a 2.91 times smaller rate constant for the

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oxidation of Ag2S (1.1×104 M-1 s-1) as obtained when assuming complete oxidation of S(-II) to

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S(+VI) in the competitive reaction kinetics experiment (see SI) would still be in the same range

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as the second order rate constant of other micropollutants. Furthermore, the Ag2S particles used

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in this study were synthesized in the absence of organics. Ag2S nanoparticles formed in the

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presence of organic compounds may be less crystalline than the particles studied here and may

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exhibit a higher sulfide oxidation rate. Based on these considerations, it is therefore reasonable to

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assume that Ag2S nanoparticles in WWTP effluent will be completely dissolved if the effluent is

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treated with ozone to remove micropollutants such as bisphenol A.

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Ozone Treatment of WWTP Effluent Spiked with Nano-Ag2S.

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In the experiments in which nano-Ag2S spiked to WWTP effluent was treated with ozone,

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Ag2S was oxidized and Ag was found in the AgNH3 fraction. However, at higher nano-Ag2S

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concentration (11.5 µM) the AgNH3 fraction increased slowly and did not reach the expected

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concentration based on complete oxidation of the spiked Ag2S (Figure 2D). Competition of

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DOM with Ag2S for the reaction with ozone may explain the delayed and slow increase of AgNH3

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in the effluent water. As ozone was additionally consumed by DOM, there was not enough ozone

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available to oxidize all nano-Ag2S even at higher specific ozone doses. In the experiments with

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less Ag2S (2.8 µM), nano-Ag2S was completely oxidized during treatment with a specific ozone

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dose of 1 gO3 / gDOC (Figure 2C). The experiments conducted in the presence of humic acid and

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absence of other organics showed that the complete oxidation of the spiked nano-Ag2S (2.8 µM)


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is immediate with a specific ozone dose of 0.24 gO3 /gDOC, suggesting that humic acid only

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marginally interfered with the oxidation of Ag2S (Figure 2A and B).

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Based on thermodynamic calculations silver sulfate (Ag2SO4) is predicted to completely

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dissolve at the applied Ag concentrations but AgCl will precipitate (Figure S5).40 LCF analysis

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of XAS spectra indicated that Ag was dominantly present in the form of Ag2S and AgCl and at

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high ozone dosages, mainly AgCl was observed suggesting that the ozone treatment resulted in

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the oxidation of Ag2S followed by precipitation of AgCl particles. The addition of NH3 to the

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experimental suspensions containing AgCl resulted in the dissolution of AgCl due to the

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formation of a soluble complex [Ag(NH3)2]+. The Ag fractions derived from the NH3 extractions

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were in good agreement with the fraction of Ag assigned to AgCl resulting from LCF analysis of

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XAS spectra, supporting the hypothesis that the NH3 extractions represent AgCl.

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The Agfree fraction found in experiments conducted at lower ozone doses most likely

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represented Ag complexed with DOM moieties, since equilibrium calculations including Ag-

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chloride complexes indicated an about eight times lower total dissolved Ag concentration (Figure

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2C). These moieties may have become oxidized at higher specific ozone doses, and, therefore,

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this fraction decreased with increasing ozone dose. In the presence of humic acid, Cl- and PO43-,

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a similar pattern was observed for the Agfree fraction, suggesting that the humic acid was to some

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extent able to complex the Ag at low ozone concentrations (Figure 2B).

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The differences in shape (rod-like vs. cube-like) and size between the Ag2S particles spiked to

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WWTP effluent and the AgCl particles formed after ozonation revealed by TEM analysis

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indicated that the transformation from Ag2S to AgCl proceeded via an oxidative dissolution -

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reprecipitation pathway.

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Effect of the Ozonation on the Toxicitity of Nano-Ag2S Spiked to WWTP Effluent.


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Pristine nano-Ag2S added to ozone treated effluent water showed no toxicity upon short-term

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exposure towards Chlamydomonas reinhardtii. Ag2S is about four orders of magnitudes less

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toxic than Ag+,4, 41 due to its very limited solubility and inertness under oxic conditions.19 These

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results indicate that nano-Ag2S does not increase the toxicity towards Chlamydomonas

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reinhardtii compared to bulk-Ag2S.

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During ozone treatment nano-Ag2S was oxidized, Ag cations were released into the

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experimental media and in the presence of Cl- precipitated as AgCl. The EC50 values from

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respective experiments were only slightly higher than the reported EC50 values for AgNO3 in

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MOPS buffer.42 Thermodynamic calculations for a simplified system (Figure S5) revealed that

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about 200 nM Ag are present as dissolved species (mainly AgCl2- and Ag+) under these

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experimental conditions. Thus, the observed toxicity may be explained by the dissolved Ag

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fraction in equilibrium with AgCl.

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Chlamydomonas reinhardtii exposed to AgNO3 spiked to (untreated) WWTP effluent resulted

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in EC50 values that were (i) twice as high as EC50 values reported for Ag+ 42 and (ii) substantially

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higher than EC50 values derived from exposure experiments conducted with nano-Ag2S spiked

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effluent water after ozonation. The reduced toxicity of Ag+ observed in this study compared to

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literature data can be explained by the formation of AgCl precipitates and the presence of

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organics in the effluent water which may complex or bind Ag cations strongly enough to further

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limit their bioavailability and therefore reduce the toxicity of Ag. Furthermore, we did not

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observe any significant differences for short-term toxicological effects towards Chlamydomonas

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reinhardtii between ozone treated and untreated effluent water.

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Environmental Implications.


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The sulfidation of Ag-NP is known to strongly decrease the short and long term toxicity of

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Ag.17 Results from this study demonstrate that during an ozone treatment of WWTP effluent at

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specific ozone doses comparable to recommended values in full-scale WWTPs, nano-Ag2S will

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be oxidized along with other micropollutants. Therefore, depending on the composition of the

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wastewater (concentration of Cl- and other Ag ligands), the short-term toxicity of the effluent

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water may increase, due to the oxidation of Ag2S and the concomitant release of Ag cations. In a

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recent study we reported Ag concentrations of a few tens of ng L-1 in the effluent of a full-scale

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WWTP.14 Under such conditions, short term toxicological effects as observed in the current

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study are unlikely, which is in agreement with a recent study reporting a consistent decrease of

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the algal toxicity after ozonation of different wastewater effluents.43 However, considerably

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higher Ag concentrations in wastewater effluent (several µg L-1)44 and in sludge (800 mg kg-1)45

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have been reported and ecotoxicological effects caused by the ozonation of such Ag-rich

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effluents cannot be completely ruled out.


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Acknowledgements

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We acknowledge the Electron Microscopy Centers at ETH Zurich (EMEZ, Zurich,

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Switzerland) and at Empa (Swiss Federal Institute for Materials Science and Technology,

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Duebendorf, Switzerland) for providing access to the microscopes. The Swiss Light Source

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(SLS, Villigen, Switzerland) is acknowledged for the allocation of beamtime. We thank Maarten

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Nachtegaal and Grigory Smolentsev (SLS) for support at the SuperXAS beamline (SLS). This

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work is part of the project “Behavior of silver nanoparticles in a wastewater treatment plant”

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within the Swiss National Research Program NRP 64 “Opportunities and Risks of

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Nanomaterials”. Additionally we would like to thank Brian Sinnet, Bettina Wagner and Elisabeth

413

Sahli for the support in the laboratories at Eawag.

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Figure 1. A: Ozone concentrations plotted against sulfate (SO42-) concentrations (error bars = 1σ

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from triplicate experiments). Dashed line represents the sulfate concentration resulting from the

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complete oxidation of the spiked nano-Ag2S (12.5 µM). The slope of the linear regression (line)

419

to the first six sulfate measurements is equal to 0.343(5) with r2 = 1.00. B: [CH2O]0/[CH2O]-1,

420

derived from measured formaldehyde concentrations, against [Ag2S]/[Buten-2-ol]. The slope of

421

the linear regression is 0.40(1) (r2 = 0.95) which corresponds to kBu/kAg2S, resulting in a rate

422

constant (kAg2S) of 3.1(1) ×104 M-1 s-1 (error bars = 1σ from experiments with the same

423

[Ag2S]/[Buten-2-ol] ratio (0.x and 0.y), but different absolute concentrations ([Ag2S]/[Buten-2-

424

ol] = 0.32: Ag2S = 16 and 48 µM; [Ag2S]/[Buten-2-ol] = 0.40: [Ag2S] = 20 and 32 µM).


22

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425 426

Figure 2. Dissolved Ag (Agfree: blue circles), NH3-soluble Ag (AgNH3: green triangles) and

427

EXAFS LCF-derived Ag2S (yellow diamonds) and AgCl (orange squares) fractions of the total

428

Ag concentrations (A, B & C: 2.8 µM, D: 11.5 µM) versus specific ozone to DOC (4 mg L-1)

429

doses. Nano-Ag2S was spiked to DDI water (buffered with 10 mM borate to pH 8.0) containing

430

humic acid (4 mg L-1, used as a proxy for DOM) in the absence (A) and presence (B) of chloride

431

(3.75 mM) and phosphate (0.1 mM) and to WWTP effluent (C & D). The right axis refers to the

432

Ag concentrations of the respective Ag fractions. The LCF fractions are scaled to the total Ag

433

concentration spiked to the WWTP effluent.


23


Page 24 of 32

434 435

Figure 3. EXAFS spectra of reference and experimental samples from Figure 2D (lines) and

436

reconstructed LCF spectra (open circles) using Ag2S and AgCl as reference spectra. For clarity

437

the spectra are plotted with an offset of 0.1. The LCF-derived fractions of Ag2S and AgCl are

438

given in Figure 2D.

439


24

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440 441

Figure 4. TEM (HAADF) image and EDX spectrum (inset) of particles collected from Ag2S-

442

spiked WWTP effluent after ozonation.

443


25


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444 445

Figure 5. Dose-response curves of the photosynthetic yield of Chlamydomonas reinhardtii after

446

1 h (filled symbols and solid line) and 2 h (open symbols and dashed line) of exposure to nano-

447

Ag2S-spiked WWTP effluent after ozonation (orange squares) and to AgNO3-spiked untreated

448

effluent water (blue circles). The dose-response curve for nano-Ag2S spiked to untreated WWTP

449

effluent is displayed in green triangles. Nano-Ag2S experiments and measurements were done in

450

triplicates, AgNO3 experiments were done as singles.

451


26

Page 27 of 32


452

Table 1. Effect Concentration (EC) Values Derived from the Photosynthetic Yield of

453

Chlamydomonas reinhardtii Determined with a Logistic Modela and Expressed as Total Ag

454

Concentration. Treatment

Time [h]

EC50 (Ag) [µM]b

Number of experiments

Ag2S and O3

1

0.256(15)

3

2

0.375(31)

3

1

0.383(4)

1

2

0.403(6)

1

AgNO3 in untreated effluent water

No short-term effect was found for Ag2S (triplicates) spiked to untreated WWTP effluent water.

455

a

Logistic model: %photosynthetic yield = min + ((max-min)/1+([Ag]total/EC50)Hillslope)

456

b

Standard error (1σ) is given in brackets


27


457

Page 28 of 32

TOC

458


28

Page 29 of 32


459

ASSOCIATED CONTENT

460

Additional supportive information is provided for the competition kinetics, characterization of the Ag2S starting material (Figure

461

S1, Table S1), measured Ag concentrations in ozonation experiments (Figure S2), Ag K-edge XANES spectra and corresponding

462

LCF results of Ag2S-spiked effluent water after ozonation (Figure S3), comparison of the Ag2S fractions derived from LCF

463

analysis of EXAFS and XANES spectra (Figure S4), chemical equilibrium concentration plot for Ag (Figure S5), characteristics

464

of nano-Ag2S (Table S1), effluent water characteristics and concentrations of selected metals (Table S2), conducted experiments

465

and applied methods (Table S3), EXAFS sum of fractions for unconstrained LCF (Table S4). This material is available free of

466

charge via the Internet at http://pubs.acs.org.


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Effect of Ozone Treatment on Nano-Sized Silver ... - ACS Publications

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