Exploring Trends of C and N Isotope Fractionation to Trace

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Exploring Trends of C and N Isotope Fractionation to Trace Transformation Reactions of Diclofenac in Natural and Engineered Systems Michael Peter Maier, Carsten Prasse, Sarah G. Pati, Sebastian Nitsche, Zhe Li, Michael Radke, Armin H. Meyer, Thomas B. Hofstetter, Thomas A. Ternes, and Martin Elsner Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b02104 • Publication Date (Web): 16 Sep 2016 Downloaded from http://pubs.acs.org on September 18, 2016

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Environmental Science & Technology

1

Exploring Trends of C and N Isotope Fractionation

2

to Trace Transformation Reactions of Diclofenac

3

in Natural and Engineered Systems

4

Michael P. Maier1, Carsten Prasse2,3, Sarah G. Pati4, Sebastian Nitsche1, Zhe Li5, Michael

5

Radke5, †, Armin Meyer1, Thomas B. Hofstetter4, Thomas A. Ternes2, and Martin Elsner1*

6

1

Helmholtz Zentrum Muenchen, German Research Center, Institute of Groundwater Ecology, Ingolstädter Landstrasse 1, Neuherberg D-85764, Germany

7 2

8

German Federal Institute of Hydrology (BfG), Referat G2, Am Mainzer Tor 1, 56068 Koblenz, Germany

9

10

3

University of California, Berkeley, Department of Civil & Environmental Engineering, Berkeley, California

11 4

12

Eawag, Swiss Federal Institute of Aquatic Science and Technology, 8600 Dübendorf,

13

Switzerland, and Institute of Biogeochemistry and Pollutant Dynamics (IBP), ETH Zürich,

14

8092 Zürich, Switzerland 5

15

University, Svante Arrhenius väg 8, SE-114 18 Stockholm

16

17 18

ACES Department of Environmental Chemistry and Analytical Science, Stockholm

†

Present address: Institute for Hygiene and Environment, Marckmannstraße 129b, 20539 Hamburg, Germany

ACS Paragon Plus Environment


19

* Corresponding author: phone: +49(0)89 3187 2565; fax: +49(0)89 3187 2565; e-mail:

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[email protected]

21

KEYWORDS. Transformation Products, Isotope Enrichment, Pharmaceuticals, Photolysis,

22

Ozonation, MnO2, ABTS, Single Electron Transfer, Micropollutants

23 24 25

ABSTRACT

26

Although diclofenac ranks among the most frequently detected pharmaceuticals in surface

27

waters, its environmental transformation reactions remain imperfectly understood.

28

Biodegradation-induced changes in

29

compound-specific isotope analysis (CSIA) may detect diclofenac degradation. This singular

30

observation warrants exploration for further transformation reactions. The present study

31

surveys carbon and nitrogen isotope fractionation in other environmental and engineered

32

transformation reactions of diclofenac. While carbon isotope fractionation was generally

33

small, observed nitrogen isotope fractionation in degradation by MnO2 (εN= -7.3‰ ± 0.3‰),

15

N/14N ratios (εN= -7.1‰ ± 0.4‰) have indicated that


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photolysis (εN= +1.9‰ ± 0.1‰) and ozonation (εN= +1.5‰ ± 0.2‰) revealed distinct trends

35

for different oxidative transformation reactions. The small, secondary isotope effect

36

associated with ozonation suggests an attack of O3 distant from the N-atom. Model reactants

37

for outer-sphere single electron transfer generated large inverse nitrogen isotope fractionation

38

(εN= +5.7‰ ± 0.3‰) ruling out this mechanism for biodegradation and transformation by

39

MnO2. In a river model, isotope fractionation-derived degradation estimates agreed well with

40

concentration mass balances, providing a proof-of-principle validation for assessing

41

micropollutant degradation in river sediment. Our study highlights the prospect of combining

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CSIA with transformation product analysis for a better assessment of transformation

43

reactions within the environmental life cycle of diclofenac.

44

45 46

INTRODUCTION Pharmaceuticals are used world-wide in large amounts. They are only partially metabolized

47

in the human body1,

2

48

(STPs). Most pharmaceuticals are incompletely eliminated in STPs

49

low-level presence (µg/L to sub-µg/L) in receiving waters with insufficiently characterized

50

effects for environmental health 4, 5. At the same time, water scarcity caused by an increasing

51

human population brings about the need of water reuse in more and more regions

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worldwide.6, 7 Direct and indirect potable reuse therefore raises the question how well natural

53

and engineered treatments are able to eliminate pharmaceuticals and other micropollutants.

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Assessments over the life cycle of a pharmaceutical in the aquatic environment – from waste

55

water to drinking water – are complicated by the difficulty of establishing mass balances in

56

natural systems such as rivers

and therefore discharged subsequently to sewage treatment plants

8, 9

2, 3

leading to a constant,

. Furthermore, abiotic and biotic transformation pathways



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are incompletely understood for most pharmaceuticals limiting our ability to assess their

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degradation by detection of transformation products (TPs).

59 60

The measurement of isotope fractionation in pharmaceuticals was recently brought forward

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as a new, complementary approach to assess their natural transformation

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antiphlogistic diclofenac – one of the most widely used12 and most frequently detected

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pharmaceuticals13, 14 – as a model compound, we accomplished accurate compound-specific

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carbon and nitrogen isotope analysis (CSIA) of diclofenac in spiked river water samples at

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concentrations down to 1 µg/L. In the same study, a significant increase of

66

15

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aerobic biodegradation and chemical reductive dehalogenation over a Pd catalyst)

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results demonstrate the potential to trace diclofenac transformation through the “footprint” of

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kinetic isotope effects of degradation reactions, just by analyzing isotope ratios in the parent

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pharmaceutical – even without detection of transformation products or knowledge about the

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degradation pathway. While analytical method development is currently aiming to decrease

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the limits of precise diclofenac isotope analysis to even lower concentrations (sub-µg/L) that

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are typically found in the environment

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investigate what insight can be obtained for assessing diclofenac transformations with the

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current method. This is particularly important, since micropollutant (e.g., pesticide and

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pharmaceutical) isotope analysis is an emerging field

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fractionation of diclofenac as a “model” pharmaceutical can, therefore, help delineating

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prospects and limitations of the general approach.

10, 11

. Taking the

13

C/12C and/or

N/14N isotope ratios in diclofenac was observed during two transformation reactions (i.e., 10

. These

2, 15

, further research is needed in parallel to

16

. Transformation-associated isotope

79 80

The potential of CSIA for tracing transformation of diclofenac in the environment, therefore,

81

warrants assessment of additional transformation reactions that have been reported for


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diclofenac in the course of its environmental life (Fig. 1). Such transformations include

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oxidations in natural and engineered systems, in particular reactions with different

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mechanisms of oxidation that would not be distinguished from TP analysis alone, because the

85

same products are formed (Fig. 1). After biological treatment – which typically cannot

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completely eliminate diclofenac in sewage treatment – diclofenac may be transformed via

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UV treatment, ozonation or react with manganese oxides in the course of additional STP

88

treatment processes.

89

released into the environment: in rivers, diclofenac can be transformed through either photo-

90

or biotransformation. 17, 20-22,10, 20, 23,24

17-19

Some of these processes are also at work when STP effluents are

91 92

Fig. 1. Transformation reactions during the environmental life cycle of diclofenac and the

93

gap of knowledge of associated isotope enrichment factors (ε); single electron oxidation

94

(SET) (2-4), photolysis20,

95

biotransformation10,

96

intermediates as discussed in more detail below. TP 4-8 were identified in this study;

28

25

(4,5), ozonation18,

26

(7-10) oxidation by MnO2-27 and

(6-8). Structures in brackets indicate hypothetical short-lived

97



98

Besides the research needed to explore trends of isotope fractionation, such CSIA data also

99

offers the opportunity to provide insight into underlying transformation mechanisms. For

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oxidative (bio)transformation of diclofenac such an independent line of evidence is

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particularly important, since it is uncertain if hydroxylated diclofenac – the most frequently

102

detected TP in previous studies (Fig. 1)

103

transformation pathway. Specifically, quantitative data showed that this TP accounted for

104

only a few percent of diclofenac transformation.10 Furthermore, the fact that hydroxyl-

105

diclofenac can be formed by different reaction pathways (Fig. 1) illustrates that the detection

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of TPs delivers only the result of transformation, but not necessarily direct insight into the

107

type of reaction that is occurring. In contrast, isotope fractionation measured by CSIA is

108

determined by isotope effects of rate-limiting steps in the underlying (bio)chemical reactions.

109

Unlike insight from transformation products (which represent the net outcome of a reaction)

110

CSIA therefore detects an indicator of the manner and order of bond breaking (i.e., transition

111

states and underlying transformation mechanisms). 31, 32

10, 28-30

– really reflects the compound’s dominant

112 113

A third prominent research gap concerns the question whether the isotope fractionation

114

measured in laboratory batch experiments – where degradation is followed under controlled

115

conditions – can also be observed in environmental systems where water parcels with a

116

different degradation history mix. It is well-recognized that differences in isotope ratios are

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levelled off in such as case so that evidence from isotope fractionation may underestimate

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degradation

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systematically addressed yet, is micropollutant degradation in rivers. If biotransformation

120

takes place in the river sediment, but samples are taken in the free water phase 24, it can be

121

expected that flowing water – where only little degradation takes place – mixes with sediment

122

pore water – which is subject to strong degradation. Whether biotransformation in rivers can

33, 34

. A scenario where this possibility is particularly relevant but has not been


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be tracked by changes in compound-specific isotope ratios under such conditions is currently

124

unclear and warrants systematic investigation.

125 126

The present study aims to investigate carbon and nitrogen isotope fractionation of important

127

environmental and engineered transformation reactions of diclofenac focusing on three

128

aspects: (i) Which transformation reactions (Fig. 1) are accompanied by isotope fractionation

129

and can thus be tracked by CSIA? To this end, isotope fractionation was studied during

130

phototransformation, transformation by manganese oxides and ozonation at ambient

131

temperature and under circumneutral pH thereby complementing previous data on aerobic

132

biodegradation10. (ii) Can CSIA shed light on different mechanisms of diclofenac oxidation?

133

To probe for an oxidative outer-sphere single electron transfer as possible mechanism of

134

aerobic degradation, diclofenac was brought to reaction with electrochemically oxidized

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ABTS (2,2'-azino-bis(3-ethylbenzothiazoline-6-sulphonic acid) radicals as model reactants.28,

136

34

137

degradation in river sediment in the same way as in groundwater? To this end, evidence from

138

diclofenac concentration and isotope analysis was evaluated at pH 8 in a circulating river

139

flume model in the dark where reactivity had previously been shown to reside in the sediment

140

35

(iii) Can we demonstrate that isotope fractionation is able to detect micropollutant

..

141 142

MATERIALS AND METHODS

143

Chemicals and General Approach. A complete list of chemicals used in this study can be

144

found in the Supporting Information (Section S1). In this explorative study we aimed to

145

conduct the experiments under realistic conditions at circumneutral pH (between 6 and 8)

146

rather than targeting a range of different pH as it may be the goal in detailed mechanistic

147

studies.



148

Phototransformation. Quartz glass tubes (39 cm long, 4.5 cm in diameter) were filled

149

with 700 mL Millipore water or river water (sampled from Isar river, Germany) spiked with

150

diclofenac (C0 = 100 mg L-1) and were exposed to sunlight from 9 a.m. to 8 p.m. on a sunny

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summer day (latitude 48.22° N). Samples were taken at different time points and split for

152

concentration and isotope analysis. For concentration and TP analysis (using LC-MS/MS)

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100 µL sample aliquots were mixed with 800 µL Millipore water and 100 µL internal

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standard (13C6-diclofenac, c = 25 mg L-1, in methanol) and filtered with a PTFE filter

155

(0.22 µm). For GC-IRMS analysis larger sample volumes were taken (10-60 mL) and

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extracted three times with dichloromethane after addition of HCl (Ctotal ~ 0.05 M). The

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dichloromethane extract was evaporated to dryness and samples were methylated by

158

BF3/methanol as described previously.10

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Ozonation. Groundwater (DOC ~ 1.5 mg L-1) was spiked with diclofenac (C0 = 30 mg L-1)

160

and different amounts of ozone to obtain O3-to-diclofenac ratios between 1:7.5 and 10:1. An

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aqueous stock solution of ozone (approx. 1 mM) was prepared by sparging ozone-containing

162

oxygen through deionized water cooled in an ice-bath. Ozone was generated from an O3-

163

generator (Ozon generator 300, Fischer Technology, Germany). To exclude the influence of

164

OH-radicals, duplicate experiments were performed in the presence of tert-butanol (100 mM)

165

as radical scavenger. Samples were prepared as described above for the phototransformation

166

experiment, except that those for GC-IRMS analysis (240 mL) were freeze-dried,

167

reconstituted in 1 mL water and subsequently liquid-liquid extracted.

168

Oxidation by MnO2. A MnO2 stock solution was synthesized according to Murray et al.36

169

and oxidation of diclofenac by MnO2 was accomplished in analogy to Forrez et al.37 Briefly,

170

900 mL Millipore water were mixed with 8 mL of NaOH (1 M) and 40 mL of NaMnO4

171

(0.1 M) under constant sparging with N2. 60 mL of a MnCl2 solution (0.1 M) were added

172

dropwise under continuous stirring. MnO2 particles were allowed to settle and the supernatant


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was replaced by de-ionized water until the electric conductivity of the solution was below

174

3 µS cm-1. This stock solution (final volume 1 L) was stored at 7 °C and used within one

175

week. Diclofenac oxidation was initiated by adding 100 mL of MnO2 stock solution to

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900 mL diclofenac solution (50 mg L-1, pH 6.2, 10 mM NaH2PO4 / Na2HPO4) under

177

continuous stirring. Because 75% of diclofenac was transformed within 10 min and the

178

solution contained only chemicals in deionised water, a biological transformation can be

179

ruled out (Supporting Information S8). Samples for concentration and isotope analysis (10-

180

60 mL) were prepared as described for the phototransformation experiment, with the

181

exception that MnO2 particles were dissolved by addition of 20% ascorbic acid (v/v, 20 mM,

182

pH 11) prior to filtration or extraction.

183

Single

electron

transfer

by

electrochemically

oxidized

2,2'-azino-bis(3-

184

ethylbenzothiazoline-6-sulphonic acid (ABTS). Oxidation of diclofenac by ABTS was

185

performed in an anoxic glovebox using anoxic stock and buffer solutions as described

186

previously.38 ABTS●- was generated by direct electrochemical oxidation of ABTS2- (0.5 mM,

187

pH 6.2, 0.1 M KH2PO4, 0.1 M NaCl) at a potential of 0.79 V (SHE) in an electrolysis cell

188

described by Aeschbacher et al.39 The working current was monitored until a stable

189

background value was reached. Variable amounts of ABTS●- (0-245 µM) were immediately

190

added to amber glass vials containing buffered diclofenac solution (30 mg L-1, pH 6.2, 0.1 M

191

KH2PO4, 0.1 M NaCl) resulting in different degrees of diclofenac oxidation in each of the

192

30 mL samples. Samples for concentration and isotope analysis (30 mL) were taken and

193

prepared as described for the phototransformation experiment, except that no HCl was added

194

during DCM extraction. This lack of HCl addition had no influence on the complete

195

extraction of diclofenac from the aqueous phase (recovery ~100%).

196

River Flume. An experiment was conducted in a bench-scale recirculating flume which

197

allows continuous transport of water relative to the sediment and thus approximates



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conditions in rivers. The experiment was conducted in the dark and in the absence of

199

precipitation. Details on flume construction and parameterization can be found in Kunkel and

200

Radke 24 and Li et al.35 Briefly, the flume was filled with 100 L water and 60 L sediment that

201

was collected from Lake Largen (north of Stockholm, Sweden), which has a negligible

202

background level of diclofenac and a pH around 8 35. After the system was equilibrated at a

203

flow velocity of 0.13 m s-1 for one week, an aqueous diclofenac solution was spiked into the

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surface water to yield an initial concentration of approximately 100 µg L-1. Surface water was

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sampled from the flume at increasing time intervals during 80 days (at hour 2, day 10, 20, 40,

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and 80 after spiking), the sampled volume was 0.3 L, 0.6 L, 1.2 L, 2.4 L, and 4.8 L,

207

respectively. Immediately after sampling, the samples were adjusted to pH 10 with a sodium

208

hydroxide (NaOH) solution (concentration: 1 M) before the samples were stored in glass

209

bottles at -18 °C until extraction. Diclofenac and 4’-OH-diclofenac concentrations were

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measured and calculated as described in Li et al.

211

described in Maier et al.

212

Waters, Eschborn, Germany) were used for extraction at neutral pH. This extraction

213

procedure had no effect on the isotopic composition of the analyte (data not shown). Due to

214

technical difficulties with the combustion reactor on the respective measurement day,

215

analysis of carbon isotope data was not successful so that only nitrogen isotope data was

216

obtained. The persistent fungicide fluconazole was used to correct for initial equilibration /

217

mixing effects by pore water and for evaporative losses 35.

218

Detection

methods.

10

35

(2015). Isotope ratios were analyzed as

with the exception that Oasis HLB cartridges (6 mL, 200 mg,

Concentrations

of

diclofenac,

4’OH-diclofenac,

and

219

phototransformation products were determined by LC-MS/MS (Agilent 1200 binary pump

220

coupled to an ABSciex API 2000 Q-TRAP mass spectrometer, see section S2 for details).

221

Samples from experiments with O3, ABTS and MnO2 were monitored for TPs using a Hybrid

222

Linear Ion Trap-Orbitrap Mass Spectrometer (LTQ Orbitrap Velos, Thermo Scientific,


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Bremen, Germany) coupled to a liquid chromatograph (Accela pump and autosampler from

224

Thermo Scientific). Full scan experiments were performed in positive electrospray ionization

225

(ESI) and atmospheric pressure chemical ionization (APCI) mode using a mass range of 60-

226

600 m/z. Data-dependent acquisition was used to obtain further information of the fragment

227

ions. To this end, full-scan experiments were followed by MS2 and MS3 scans for the two

228

most intense ions. An external mass calibration was performed prior to the analysis of each

229

batch to ensure accurate mass determinations with a resolution of (m/z)/∆(m/z) = 60,000. A

230

mixture of n-butylamine, caffeine, and Ultramark 1621 (mixture of fluorinated phosphazines)

231

was used for mass calibration. The mass accuracy was always within 0.5 ppm.40

232

Isotope analysis. Isotope ratios of diclofenac were analyzed by GC-IRMS as described

233

previously.10 Briefly, methylated samples (in hexane) were either injected with a split ratio of

234

1:10 or splitless into a split/splitless injector (Thermo Fisher Scientific) at 280 °C with a flow

235

rate of 1.4 mL min-1. Separation was achieved by a gas chromatograph (TRACE GC Ultra

236

gas chromatograph, Thermo Fisher Scientific) equipped with a DB-5 column (30 m ×

237

0.25 mm, 1 µm film thickness, J&W Scientific, Folsom, Canada). The GC temperature was

238

ramped from 80 °C (1 min) to 200 °C with a rate of 17 °C min-1 and then at 6 °C min-1 to

239

300 °C (held for 2 min). After chromatographic separation diclofenac was combusted in a

240

Finnigan GC combustion interface (Thermo Fisher Scientific) to CO2 and N2 with a NiO tube

241

/ CuO-NiO reactor operated at 1000 °C (Thermo Fisher Scientific). Isotope values of CO2

242

and N2 were determined with a Finnigan MAT 253 isotope ratio mass spectrometer (Thermo

243

Fisher Scientific). For quality control, a diclofenac lab standard with known isotopic

244

signature was analyzed in the same way as the samples at least every ninth injection. All

245

reported isotope values were corrected to international reference materials, Vienna PeeDee

246

Belemnite for carbon (Eq. 2) and Reference Air for nitrogen isotopes (Eq. 3).41 δ13C values

247

were additionally corrected for the introduced methyl-group.10



= =

/ − /

/

!

! "

/ # − / #

/ #

!

Page 12 of 32

(1)

! "

(2)

248

Equation 3 (“Rayleigh equation”)42 was used to evaluate how isotope values of a

249

micropollutant at different points in time (δ13Ct, δ15Nt, respectively) changed relative to the

250

beginning of a transformation (δ13C0, δ15N0, respectively) in dependence of the extent of

251

transformation (Ct/C0). This equation was used to obtain reaction-specific enrichment factors

252

ε by fitting experimental data in SigmaPlot™ (here expressed for δ15N)

253 254

=

1 + 1 +

(3)

If ε is known, equation 4 can be used in field situations to calculate how much diclofenac is not yet transformed (f = Ct/C0) using measured isotope values (here expressed for δ15N). 43 /

1 + = =

1 +

(4)

255 256

RESULTS AND DISCUSSION

257

(i) Exploring trends and magnitude of isotope fractionation in environmental and

258

engineered transformation reactions of diclofenac

259

Oxidation by MnO2. Diclofenac was transformed by MnO2 and hydroxylated diclofenac

260

as well as its corresponding quinone were detected as TPs (Fig. 1, Fig. S5, Fig. S8).37 Small,

261

but significant C-isotope fractionation (εC = -1.5 ± 0.1‰) as well as pronounced N-isotope

262

fractionation (εN = -7.3 ± 0.3‰) was observed during diclofenac oxidation by MnO2 (Fig. 2)

263

For both elements, isotope fractionation occurred in the normal direction meaning that

264

diclofenac molecules containing light isotopes (12C,

265

heavy isotopes (13C,

15

14

N) reacted faster and molecules with

N) were left behind in the residual substrate fraction so that


13

C/12C

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15

N/14N ratios increased over time. Hence, diclofenac oxidation by MnO2 – which can

266

and

267

either occur during water treatment where manganese oxide is used as a flocculant

268

soils and sediments where it is naturally present 44 – can be detected by CSIA.

19

or in

269 270

Fig. 2. Carbon (left) and nitrogen (right) isotope fractionation of diclofenac during

271

transformation in engineered systems: by MnO2 (blue data points) and during ozonation (red

272

data points). Error bars correspond to typical uncertainties associated with the respective

273

isotope analysis (±1‰ for δ15N analysis, ±0.5‰ for δ13C analysis). Regressions according to

274

equation 3 are graphically represented together with their 95% confidence intervals

275

corresponding to the uncertainties of the reported enrichment factors ε (Table 1).

276

Oxidation by ozone. In contrast to oxidation by MnO2, ozonation of diclofenac caused an

277

15

278

inverse N-isotope fractionation meaning that

N containing molecules reacted faster and

279

15

280

diclofenac-2,5-iminoquinone were detected (Fig. 1 (8), Fig. S5).18,

281

experiment ozone was applied in the presence of the radical scavenger tert-butanol to quench

282

hydroxyl radicals.26 Smaller O3 doses were needed to transform diclofenac under these

283

circumstances, since less O3 was lost in side reactions with OH● radicals (see discussion in

284

Supporting Information S6). Isotope fractionation, however, showed no significant difference

N/14N ratios decreased (εN = +1.5 ± 0.2‰, Fig. 2). Simultaneously, oxidized TPs such as


45, 46

In a second


285

between experiments with and without the radical scavenger (Tab. 1) indicating that

286

transformation – and isotope fractionation – was solely attributable to the reaction of

287

diclofenac with ozone but not with OH●. This is consistent with expectations from reported

288

rate constants (see detailed discussion in the Supporting Information S6). Due to the small

289

but distinct εN-value, CSIA can be used to detect transformation by ozonation if changes in

290

δ15N values are twice the analytical uncertainty of ± 1‰ corresponding to an extent of

291

degradation above 70% (i.e., f ≤ 0.3). These findings may not only be interesting for

292

diclofenac, but potentially also for other micropollutants containing aromatic amine

293

structures.

294

Photolysis. C and N isotope ratios of diclofenac were analyzed during photolysis under

295

natural sunlight in river water. Normal carbon and inverse nitrogen isotope fractionation was

296

observed (εC = -0.7 ± 0.2‰, εN = +1.9 ± 0.4‰) (Fig. S10). Similar isotope fractionation in

297

ultrapure water (εC = -1.1 ± 0.4‰, εN = +2.0 ± 0.2‰) (Fig. S10) confirmed the hypothesis

298

that direct rather than indirect photolysis caused diclofenac transformation 47. Consequently,

299

isotope effects appear to be robust towards varying organic matter or salt compositions at

300

circumneutral pH, at least in the range of the two tested water samples (Tab. S4).

301

Figure 3 compares the photolysis-induced isotope fractionation (combined results in river

302

and MilliQ water) to isotope fractionation observed in batch experiments for

303

biotransformation in river sediment

304

(normal for biodegradation, inverse for direct photolysis) illustrates the potential to identify

305

either process if it strongly dominates diclofenac transformation. This possibility could be

306

helpful to address the question to which extent photolysis plays a role in turbid rivers.22 Also,

307

the result confirms the qualitative observation of previous studies that inverse N isotope

308

effects are more frequent in direct photolysis due to mass independent (e.g., magnetic)

309

isotope effects48,

10

. The difference in nitrogen isotope fractionation

49

, whereas inverse N isotope effects in “conventional” (i.e., thermally


Page 14 of 32

Page 15 of 32


50

310

induced) (bio)chemical reactions

are less frequent since they occur when biological and

311

abiotic oxidations involve tighter bonds to N in the transition state.

312

313 314

Fig. 3. Carbon (left) and nitrogen (right) isotope fractionation of diclofenac in surface water

315

and sediments: during photolysis in river water (this study, blues data points) and during

316

biotransformation in river water/sediment (adopted from Maier et al.,10 red data points). Error

317

bars correspond to typical uncertainties associated with the respective isotope analysis (±1‰

318

for δ15N analysis, ±0.5‰ for δ13C analysis). Regressions according to equation 3 are

319

graphically represented together with their 95% confidence intervals corresponding to the

320

uncertainties of the reported enrichment factors ε (Table 1).

321

(ii) Exploring the usefulness of observed isotope effects to provide insight into

322

underlying molecular mechanisms of different oxidative transformations.

323

The contrasting trends in nitrogen isotope fractionation during different transformation

324

reactions of diclofenac may further be explored for their usefulness to provide insight into

325

underlying reaction mechanisms. Two observations are particularly intriguing. On the one

326

hand isotope fractionation during abiotic oxidation by MnO2 agrees almost exactly with the

327

N-isotope enrichment of putative biodegradation of diclofenac in river sediment from our

328

previous study (εN = -7.1‰)10 and the flume study discussed in detail below. On the other



329

hand isotope fractionation showed the opposite trend during ozonation, even though similar

330

TPs were observed (Fig. 1 and Fig. S5).

331

Is there a common mechanism behind oxidative transformation? - Testing for outer

332

Sphere Single Electron Transfer (SET) with ABTS. Oxidative SET is frequently proposed

333

as a putative reaction mechanism in aerobic biodegradation and oxidation by MnO2.51-53, 28, 34

334

Hence, we electrochemically generated ABTS radicals - a putative outer-sphere SET

335

reagent54 - to transform diclofenac and test the hypothesis that isotope fractionation is similar

336

in the systems mentioned above and in the ABTS model system.55 Transformation caused no

337

changes in carbon isotope ratios, but pronounced inverse N-isotope fractionation (εN = +5.7

338

± 0.3‰, Fig. 4). This was a surprise because, according to the hypothesis of SET for

339

biodegradation and oxidation by MnO2 (see above) the opposite trend – a consistent normal

340

isotope effect for SET - would have been expected. The existence of a different mechanism,

341

however, was also confirmed by detection of different TPs. During oxidation by ABTS,

342

radical coupling TPs were detected which are the hallmark of an initial one electron

343

abstraction at the N-atom because they indicate that radical cations are formed as

344

intermediates (Fig. 1 (4), Fig. S7).38 This mechanism is further supported by calculations

345

predicting an inverse N isotope effect for this specific type of SET: an outer sphere one-

346

electron abstraction leading to isolated radicals as opposed to an inner sphere SET where

347

electrons are transferred through chemical bonds.55,

348

oxidation with ABTS probably occurs by outer sphere SET. In contrast, the observed

349

differences in isotope effects of biotransformation and oxidation by MnO2 imply that these

350

reactions have other rate-limiting steps and follow a different reaction mechanism, most

351

likely inner sphere electron transfer. The direction and magnitude of isotope fractionation in

352

this putative inner sphere SET agrees well with recent results on diphenylamine oxidation

353

(εC = -2.3‰, εN = -10.0‰)57 suggesting that this may potentially represent a general pattern

56

Hence, our results indicate that


Page 16 of 32

Page 17 of 32


354

of N isotope fractionation in many oxidative natural transformations. Further studies with

355

other compounds will be needed to substantiate this pattern.

356

It appears unlikely, finally, that the formation of the small amounts of mono-hydroxylated

357

diclofenac (Fig. S8) that were detected during biotic and MnO2-facilitated transformation of

358

diclofenac

359

hydroxylation occurs in a molecular position that is distant from the C-N bond (secondary

360

isotope effect). This conclusion is supported by recent results for benzotriazole that was

361

exclusively transformed by hydroxylation at positions distant from the C-N bond and with a

362

maximum isotope enrichment of εN = -1.1‰.58 We therefore hypothesize that diclofenac is,

363

to its major part, transformed by an unknown oxidation pathway involving oxidation of the

364

nitrogen atom. Although the corresponding TPs have not been detected so far, a reaction at

365

the N-atom is indicated by the observable pronounced N-isotope fractionation.

10

caused the pronounced N-isotope fractionation of εN = -7‰, because the

366

At what molecular position occurs the attack during ozonation? – Considering the

367

magnitude of N isotope fractionation. In a similar way as for oxidation by putative SET,

368

the usefulness of isotope fractionation to provide mechanistic information can also be

369

explored for ozonation. Two sites have been suggested for the initial attack of ozone in

370

diclofenac. Comparing measured isotope ratios with mechanistic scenarios offers the

371

possibility to explore the plausibility of either scenario. An attack of ozone at the N-atom

372

would lead to a cationic intermediate (Fig. 1 (9)).

373

position of the non-chlorinated aromatic ring

374

the C-N bond as a partial imine bond as shown in compound (10) in Fig. 1.38 In both cases,

375

(i.e., for a more crowded coordination environment at the N-atom as in Fig. 1(9), or for an

376

increased double bond character as in Fig. 1(10)), the energy of molecular vibrations

377

involving the N atom would be increased in the transition state which usually leads to an

378

inverse isotope effect.18, 26, 59 However, a direct attack at the N-atom would probably cause a

18

26

Alternatively, an attack at the para-

would increase the double bond character of



379

primary N-isotope effect of greater magnitude. Therefore, a secondary isotope effect caused

380

by attack at the distant para position (Fig. 1 (10)) appears to be more consistent with the

381

observed small inverse nitrogen isotope fractionation. This scenario is corroborated by a

382

similar pathway of aniline ozonation which has been inferred from detected TPs.18

383 384

Fig. 4. Carbon (left) and nitrogen (right) isotope fractionation of diclofenac by the putative

385

Outer Sphere (OS)-SET model reagent ABTS (red symbols) to test for SET as mechanism for

386

transformation by MnO2 (blue symbols) and biotransformation (white symbols). Error bars

387

correspond to typical uncertainties associated with the respective isotope analysis (±1‰ for

388

δ15N analysis, ±0.5‰ for δ13C analysis). Regressions according to equation 3 are graphically

389

represented together with their 95% confidence intervals corresponding to the uncertainties of

390

the reported enrichment factors ε (Table 1).

391

(iii) Exploring the ability of isotope fractionation to trace degradation under

392

environmentally relevant conditions (river model): comparison of results from

393

concentration and isotope analysis. Recent experiments in a well-mixed sediment-water

394

batch system – where pore water concentrations were in equilibrium with the free water

395

compartment – suggested that nitrogen isotope fractionation is a promising indicator of

396

(bio)transformation (Figure 3).10 To investigate if this also applies to conditions where

397

rapidly flowing water interacts with underlying sediment, an experiment was conducted in a

398

recirculating river model24. No diclofenac attenuation was observed in sterile (water +


Page 18 of 32

Page 19 of 32


399

sediment + sodium azide) and water-only controls, whereas diclofenac concentrations

400

decreased in the actual experiment. This suggests that biotransformation occurred and that the

401

sediment was the reactive compartment. With the coarse sand and the low streaming velocity

402

chosen, we further observed no suspension or transport of sediment. Finally, since aerobic

403

biodegradation is strongly preferred to anaerobic degradation

404

cm of the sediment were oxic (see oxygen profile in the Supporting Information S11), we

405

conclude that aerobic biodegradation occurred in this upper part of the sediment. This

406

transformation of diclofenac was associated with pronounced normal nitrogen isotope

407

fractionation (Fig. 5), which is in very good agreement with the non-flow batch experiment,

408

even though sediment from a different river was used.10 We used this information to calculate

409

the extent of transformation using measured isotope ratios and the enrichment factor from the

410

batch experiment (Eq. 4, Fig. 5). The estimated extent of transformation derived from

411

measured diclofenac concentrations (red bars in Figure 5) can be compared to the estimate

412

from isotope ratios (blue bars in Figure 5). A distinct difference becomes apparent between

413

day 0 and day 10 when concentration measurements in the water phase indicate a stronger

414

mass elimination of diclofenac (67% of substance remaining) than calculations from isotope

415

ratios according to Equation 4 (92% remaining). The difference between the two methods is

416

25% at day 10 and corresponds to the pore water content of the sediment, which is

417

approximately 22% of the total water volume

418

penetrated into the sediment, whereas (non-spiked) pore water came out. Due to this initial

419

mixing / equilibration the extent of transformation in the initial phase was overestimated

420

when based on the decrease in measured concentrations in the surface water. In contrast,

421

information from isotope ratios was not affected by this artifact which highlights the added

422

value of CSIA as a complementary approach. In contrast to the initial phase, estimates of

423

degradation between day 10 and 20 and between day 20 and 40 derived from concentration

35

29

, and since only the upper 2

indicating that the (spiked) surface water



Page 20 of 32

424

analysis and from CSIA data agreed well and indicate approximately the same loss of

425

diclofenac by reaction. This agreement also reinforces the confidence in the nitrogen isotope

426

enrichment factor of εN ≈ -7‰ that we used. We derived this value in our previous study

427

with oxic river sediment from a very different geographic location (Isar river, Bavaria). To

428

our knowledge, no microbial diclofenac degrader strains have been isolated yet so that the

429

determination of εN values in a more defined experimental system are presently elusive. We

430

note, however, that besides Maier et al. 2014 (oxic Isar river sediment, 10) and Schürner et al.

431

2016 (oxic aquifer sediment from Bavaria in an indoor aquifer 60) this experiment in Sweden

432

is already the third study consistent with an εN of about -7‰. We, therefore, tentatively

433

conclude that these enrichment factors, even though derived under little defined conditions,

434

appear to be quite representative of aerobic diclofenac biodegradation in the environment.

435 436


10

Page 21 of 32


437 438

Fig. 5. Upper panel: nitrogen isotope isotope ratios (blue squares), concentrations of

439

diclofenac (red squares) and concentrations of 4-hydroxy diclofenac (red hollow squares)

440

during transformation in the river flume system. Lower panel: remaining diclofenac in the

441

flume estimated (i) from nitrogen isotope fractionation according to Equation 4 (blue bars)

442

and (ii) from concentration measurements of diclofenac in the water body (red bars); error

443

bars indicate the typical measurement uncertainty for δ15N of about 1 ‰).

444

ENVIRONMENTAL SIGNIFICANCE

445

As summarized in Table 1, this study reports C and N isotope fractionation for major

446

natural and engineered transformation reactions of diclofenac. These results allow assessing

447

CSIA for its ability to trace different transformation processes of diclofenac in the

448

environment and during treatment. A general trend towards normal

449

fractionation

13

C/12C isotope

suggests that if carbon isotope ratios are observed to change towards greater



Page 22 of 32

450

13

C/12C values, this line of evidence may serve as a general indicator of diclofenac

451

transformation. However, observable isotope effects were small implying that transformation

452

reactions must undergo a high degree of conversion in order to be detected by changes in

453

carbon isotope values. Observed N-isotope fractionation, in contrast, was not only larger and

454

represented, therefore, a more robust measure of transformation. N-isotope fractionation was

455

also strongly characteristic for different transformation reactions, because distinct isotope

456

fractionation could be observed in different processes (normal nitrogen isotope effects in

457

biodegradation and transformation by MnO2 vs inverse isotope effects in ozonation and

458

photolysis). A study of diclofenac transformation in a river model confirmed that such

459

pronounced nitrogen isotope fractionation can also be observed for diclofenac degradation at

460

microgram per liter concentrations in river sediments, providing an encouraging proof-of-

461

principle validation under realistic hydrogeological conditions.

462 463

Table 1. Carbon and nitrogen enrichments factors (εC and εN) for diclofenac transformation

464

in different model systems; ε-values are stated together with their standard errors.

System

εC (‰)

εN (‰)

Photodegradation in river water

-0.7 ± 0.2

+1.9 ± 0.4

Ozonation

n.s.

+1.5 ± 0.4

Manganese oxides (MnO2)

-1.5 ± 0.3

-7.3 ± 0.6

Biodegradation*

n.s.

-7.1 ± 0.9

Direct photolysis in ultrapure water

-1.1 ± 0.4

+2.0 ± 0.2

Ozonation, radical scavenger added

n.s.

+1.9 ± 0.4

Electrochemically oxidized ABTS

n.s.

+5.7 ± 0.6

Pd-catalyzed reductive dechlorination*

-2.0 ± 0.4

n.s.

Engineered & Environmental systems

Model Reactant Systems


Page 23 of 32


Combined Data

465

Ozonation

n.s.

+1.7 ± 0.3

Photolysis

-0.8 ± 0.2

+1.9 ± 0.3

*

from Maier et al.10; n.s. = not significant;

466 467

While results from this study, therefore, illustrate the prospect of environmental assessments

468

by CSIA, they also indicate two kinds of potential limitations and possible ways to overcome

469

them.

470

First, although evidence from nitrogen isotope fractionation is most promising, sensitive

471

nitrogen CSIA is particularly challenging. The reason is that diclofenac contains only one N-

472

atom, whereas two N-atoms are needed to form one molecule of N2 gas for IRMS analysis,

473

and on top the relative natural abundance of

474

Further targeted analytical development is, therefore, ongoing to accomplish sensitive isotope

475

analysis in the low concentration range (low µg/L to sub-µg/L) typically encountered in the

476

environment 16, 62, 63. The results of the present study demonstrate that this effort is, however,

477

worthwhile. Thus, our results are an important step forward towards environmental

478

assessment of micropollutants by CSIA, even though they were obtained at relatively high

479

concentrations.

15

N (0.37%) is lower than of

13

C (1.1%)

61

.

480

Second, our results suggest that evidence from nitrogen CSIA may be inconclusive if bio-

481

and photodegradation occur simultaneously because the respective trends cancel each other

482

out. For such a situation our study illustrates the added value of a combined approach. Where

483

evidence from isotope analysis would be inconclusive, identification of phototransformation

484

can be possible by detection of specific dechlorinated TPs64 (Fig. 1 (5), Fig. S3) that are not

485

observed in biodegradation

486

may indicate oxidative (bio-)degradation when TPs are inconclusive.

29, 30

. Vice versa, strong normal nitrogen isotope fractionation



487

Finally, these considerations also illustrate how a combined approach of CSIA and TP can

488

help tackling a prominent knowledge gap for many micropollutants: in contrast to legacy

489

pollutants from contaminated sites, transformation pathways of many micropollutants are yet

490

to be explored, and for many of them, pure degrader strains have not yet been isolated. TPs

491

can indicate the occurring transformation pathways and capture the presence of toxic or

492

persistent TPs. In situations, however, where the mass balance between reactant and TP(s) is

493

not closed and it is unclear if the most relevant TP(s) are detected, CSIA can provide an

494

independent line of evidence. CSIA shows if the detected TPs and observed isotope

495

fractionation are in agreement for a specific transformation pathway. In addition, when

496

important transformation pathways are not captured in TP analysis, CSIA can reveal the

497

elements involved in rate limiting reaction steps and thereby give a starting point to explore

498

unknown degradation pathways. This line of evidence may even become stronger in future

499

assessments if isotope analysis of additional elements such as chlorine becomes accessible in

500

diclofenac.

501 502

ACKNOWLEDGEMENTS

503

We acknowledge Jan Funke for assistance with the ozonation experiments and Uwe Kunkel

504

for his help during TP search. Michael Maier was financially supported by the German

505

Federal Environmental Foundation (DBU).

506

Supporting information available

507

S1. Materials; S2. LC-MS/MS analysis; S3. Detection of TPs during Phototransformation;

508

S4. Experimental Conditions of photolysis experiment; S5.

Formation of diclofenac-

509

2,5-iminoquinone in the presence of O3, ABTS or MnO2; S6.

Discussion why reaction of

510

diclofenac with hydroxyl radicals is negliglible compared to reaction with ozone; S7.

511

Detection of coupling TPs during transformation by ABTS; S8. Detection of OH-diclofenac


Page 24 of 32

Page 25 of 32


512

during MnO2 transformation; S9. Transformation of diclofenac in the river flume; S10.

513

Isotope fractionation during photolysis of diclofenac in river and MilliQ water; S11. Oxygen

514

profile in the sediment of the flume experiment;

515

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Exploring Trends of C and N Isotope Fractionation to Trace

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