Ammonia Monooxygenase-Mediated Cometabolic Biotransformation

Subscriber access provided by University of South Dakota

Environmental Processes

Ammonia monooxygenase-mediated cometabolic biotransformation and hydroxylamine-mediated abiotic transformation of micropollutants in an AOB/NOB co-culture Yaochun Yu, Ping Han, Li-Jun Zhou, Zhong Li, Michael Wagner, and Yujie Men Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b02801 • Publication Date (Web): 13 Jul 2018 Downloaded from http://pubs.acs.org on July 14, 2018

Just Accepted “Just Accepted” manuscripts have been peer-reviewed and accepted for publication. They are posted online prior to technical editing, formatting for publication and author proofing. The American Chemical Society provides “Just Accepted” as a service to the research community to expedite the dissemination of scientific material as soon as possible after acceptance. “Just Accepted” manuscripts appear in full in PDF format accompanied by an HTML abstract. “Just Accepted” manuscripts have been fully peer reviewed, but should not be considered the official version of record. They are citable by the Digital Object Identifier (DOI®). “Just Accepted” is an optional service offered to authors. Therefore, the “Just Accepted” Web site may not include all articles that will be published in the journal. After a manuscript is technically edited and formatted, it will be removed from the “Just Accepted” Web site and published as an ASAP article. Note that technical editing may introduce minor changes to the manuscript text and/or graphics which could affect content, and all legal disclaimers and ethical guidelines that apply to the journal pertain. ACS cannot be held responsible for errors or consequences arising from the use of information contained in these “Just Accepted” manuscripts.

is published by the American Chemical Society. 1155 Sixteenth Street N.W., Washington, DC 20036 Published by American Chemical Society. Copyright © American Chemical Society. However, no copyright claim is made to original U.S. Government works, or works produced by employees of any Commonwealth realm Crown government in the course of their duties.

Page 1 of 39

Environmental Science & Technology

Ammonia monooxygenase-mediated cometabolic biotransformation and hydroxylamine-mediated abiotic transformation of micropollutants in an AOB/NOB co-culture Yaochun Yu1, †, Ping Han2, †, Li-Jun Zhou2, 3, Zhong Li4, Michael Wagner2, Yujie Men1, 5, *

1

Department of Civil and Environmental Engineering, University of Illinois at

Urbana-Champaign, Urbana, IL 61801, USA 2

Department of Microbiology and Ecosystem Science, Division of Microbial Ecology,

Research Network “Chemistry meets Microbiology”, University of Vienna, 1090 Vienna, Austria 3

State Key Laboratory of Lake Science and Environment, Nanjing Institute of Geography and

Limnology, Chinese Academy of Sciences, Nanjing 210008, China 4

Metabolomics Center, University of Illinois, Urbana, IL 61801, USA

5

Institute for Genomic Biology, University of Illinois at Urbana-Champaign, Urbana, IL

61801, USA †

Equal contribution

* Corresponding Author Yujie Men Department of Civil and Environmental Engineering University of Illinois at Urbana-Champaign, Address: 205 North Mathews Ave Urbana, IL 61801-2352, USA Email: [email protected]; Phone: (217) 244-8259

ACS Paragon Plus Environment


1

Abstract

2

Biotransformation of various micropollutants (MPs) has been found to be positively

3

correlated with nitrification in activated sludge communities. To further elucidate the roles

4

played by ammonia-oxidizing bacteria (AOB) and nitrite-oxidizing bacteria (NOB), we

5

investigated the biotransformation capabilities of an NOB pure culture (Nitrobacter sp.), and

6

an AOB (Nitrosomonas europaea)/NOB (Nitrobacter sp.) co-culture for fifteen MPs, whose

7

biotransformation was reported previously to be associated with nitrification. The NOB pure

8

culture did not biotransform any investigated MP, whereas the AOB/NOB co-culture was

9

capable of biotransforming six MPs (i.e. asulam, bezafibrate, fenhexamid, furosemide,

10

indomethacin, and rufinamide). Transformation products (TPs) were identified and tentative

11

structures were proposed. Inhibition studies with octyne, an ammonia monooxygenase (AMO)

12

inhibitor, suggested that AMO was the responsible enzyme for MP transformation that

13

occurred co-metabolically. For the first time, hydroxylamine, a key intermediate of all

14

aerobic ammonia oxidizers, was found to react with several MPs at concentrations typically

15

occurring in AOB batch cultures. All of these MPs were also biotransformed by the

16

AOB/NOB co-culture. Moreover, the same asulam TPs were detected in both

17

biotransformation and hydroxylamine-treated abiotic transformation experiments, whereas

18

rufinamide TPs formed from biological transformation were not detected during

19

hydroxylamine-mediated abiotic transformation, which was consistent with the inability of

20

rufinamide abiotic transformation by hydroxylamine. Thus, in addition to cometabolism

21

likely carried out by AMO, an abiotic transformation route indirectly mediated by AMO

22

might also contribute to MP biotransformation by AOB.

ACS Paragon Plus Environment

Page 2 of 39

Page 3 of 39


23

INTRODUCTION

24

In recent years, with an increasing use of organic chemicals such as pesticides,

25

pharmaceuticals and personal care products, there are emerging concerns about the

26

potential risk of these contaminants to natural environments and public health.1-4 These

27

compounds are typically detected at very low concentrations in aquatic systems (i.e., ng/L

28

~ µg/L levels),5, 6 thus are also referred to as micropollutants (MPs). Conventional

29

wastewater treatment plants (WWTPs) are one major sink of MPs.7 Biotransformation is an

30

essential mitigation route of MPs in WWTPs, but most of these systems are not designed

31

specifically for MP removal. Thus, MP removal in WWTPs is generally incomplete and its

32

efficiency is highly compound specific.8-10 Remaining MPs and their transformation

33

products (TPs) are discharged into aquatic systems, which may cause adverse effects.11-13

34

Hence, further studies of MP biotransformation mechanisms and identification of their TPs

35

are necessary and will in the long run help us to better understand the environmental fate of

36

MPs, design removal or monitoring strategies, as well as establish precaution guidelines.2, 5,

37

14

38

Nitrification, the aerobic oxidation of ammonia via nitrite to nitrate, is an essential

39

step for the removal of nitrogen compounds in WWTPs. In many WWTPs, this process is

40

catalyzed by the concerted activity of ammonia-oxidizing bacteria (AOB) converting

41

ammonia to nitrite, and nitrite-oxidizing bacteria (NOB) converting nitrite to nitrate,

42

although ammonia-oxidizing archaea (AOA)15-18 the archaeal counterpart of AOB and the

43

recently discovered complete ammonia oxidizers (comammox) converting ammonia to

44

nitrate19-21 have also been detected in some WWTPs. Previous studies have shown 3 ACS Paragon Plus Environment


45

correlations between nitrification and biotransformation of some MPs in nitrifying

46

activated sludge communities.8, 11, 22-24 For example, MP biotransformation was enhanced

47

when nitrification activity was stimulated.25 In addition, MP biotransformation in activated

48

sludge communities decreased when nitrification was completely inhibited by ammonia

49

monooxygenase (AMO) inhibitors.12, 25-27 In a recent study, the biotransformation of a

50

number of MPs was inhibited by AMO inhibitors, and more than 50% inhibition was

51

observed for four MPs (i.e., asulam, trimethoprim, monuron, and clomazone), providing

52

strong indications that ammonia oxidizers (particularly AOB) were involved in the

53

biotransformation of some MPs.8 However, the correlations observed in inhibition studies

54

does not necessarily lead to causal relationships. In order to demonstrate the actual role

55

played by AOB in MP biotransformation, and to identify the responsible enzymes and

56

transformation pathways, complementary pure culture studies are required.

57

It has been hypothesized that AMO, a substrate-promiscuous key enzyme of all

58

autotrophic ammonia oxidizers is responsible for nitrifier-mediated MP

59

biotransformation.25, 28, 29 So far, whole-cell studies have been used to investigate AMO

60

functionality due to the unavailability of purified AMO with retained activity.30 While it has

61

been shown that AMO is able to oxidize some hydrocarbons and halogenated hydrocarbons

62

co-metabolically,31, 32 direct evidence for MP transformation by this enzyme is largely

63

lacking. Furthermore, the results of the inhibition studies could also suggest that the MP

64

biotransformation was carried out by microorganisms whose growth is highly dependent on

65

ammonia oxidizers. For example, the inhibition of ammonia oxidizers in nitrifying

66

activated sludge will strongly reduce nitrite production, and thus also inhibit NOB. 4 ACS Paragon Plus Environment

Page 4 of 39

Page 5 of 39


67

Therefore, there is an urgent need to investigate the MP biotransformation potential of

68

nitrite oxidizers, too. Given that there is no nitrite accumulation in nitrifying activated

69

sludge communities and the accumulated nitrite in AOB pure cultures may lead to abiotic

70

nitration reactions of some MPs,33, 34 a co-culture system containing AOB and NOB species

71

is more relevant to the actual nitrification condition, and will be used in this study to

72

investigate the roles of nitrifiers.

73

The goal of this study was (i) to investigate biotransformation capabilities of AOB

74

and NOB using the model AOB species Nitrosomonas europaea and NOB species

75

Nitrobacter sp.. Fifteen MPs including pharmaceuticals and pesticides were selected,

76

whose biotransformation has been associated with nitrification in activated sludge by an

77

inhibition study,8 (ii) to identify responsible enzymes for the transformation of selected

78

MPs, and (iii) to further understand the respective transformation mechanisms.

79

Biotransformation experiments were conducted using the NOB pure culture and a nitrifying

80

co-culture containing the AOB and the NOB species. Metabolic and co-metabolic

81

biotransformation mechanisms were tested by using the biotransformed MP as the sole

82

energy source. AMO inhibition experiments were carried out to demonstrate the

83

involvement of AMO in MP biotransformation. Abiotic transformation experiments using

84

the key ammonia oxidation intermediates (e.g., hydroxylamine) were also conducted to

85

further unravel mechanisms of MP biotransformation by ammonia oxidizers.

86

MATERIALS AND METHODS

87

MP Selection. Based on a previous inhibition study of MP biotransformation in activated

88

sludge, 15 MPs (Table S1) were selected, whose biotransformation was inhibited by more 5 ACS Paragon Plus Environment


89

than 20% when adding the bacterial AMO inhibitor octyne.8 These MPs are commonly

90

used pesticides and pharmaceuticals, including acetamiprid, asulam, bezafibrate,

91

carbendazim, clomazone, fenhexamid, furosemide, indomethacin, irgarol, levetiracetam,

92

monuron, rufinamide, tebufenozide, thiacloprid, and trimethoprim. In addition, two

93

compounds (mianserin and ranitidine) that were previously shown to be biotransformed by

94

ammonia-oxidizing pure cultures,5 were included as positive controls for biomass activities

95

(Table S1). Reference compounds and methanol (HPLC grade) were purchased from

96

Sigma-Aldrich (St. Louis, MO). Five internal standards including trimethoprim-d3,

97

metoprolol acid-d5, carbendazim-d4, and furosemide-d5 were obtained from C/D/N

98

isotopes (Quebec, Canada) and Toronto Research Chemicals (Toronto, Canada). The

99

bacterial AMO inhibitor 1-octyne35, 36 was purchased from Fisher Scientific. Stock

100

solutions of MPs and internal standards were prepared individually in methanol at 1 g L-1,

101

except for carbendazim and its corresponding internal standard carbendazim-d4, which

102

were prepared in methanol at 0.1 g L-1. All stock solutions were stored at -20 °C. A mixture

103

of standards (100 mg L-1 for each MP) was prepared via appropriate dilution of the stock

104

solutions in methanol.

105

Cultivation of Nitrifying Pure and Co-cultures. The AOB strain Nitrosomonas europaea

106

Winogradsky (ATCC 19718) and the NOB stain Nitrobacter sp. (ATCC 25381) were

107

obtained from the American Type Culture Collection (ATCC). They were maintained in 50

108

mL suspension culture flasks with filter screw caps (VWR, Radnor, PA) containing 20 mL

109

of a modified basal medium with 4.0 g L-1 CaCO3 for buffering the pH around 8.37 All

110

cultures were incubated at 28 °C in a dark incubator and shaken at 120 rpm. Seven mM 6 ACS Paragon Plus Environment

Page 6 of 39

Page 7 of 39


111

NH4Cl or NaNO2 was added as the growth substrate in the AOB or NOB cultures, which

112

was re-amended upon depletion. The culture performance (ammonia or nitrite oxidation)

113

was consistent and stable during each feeding cycle of the entire incubation period. The

114

AOB/NOB co-culture was obtained by growing the two strains together, and maintained in

115

the same basal medium with 2 mM NH4Cl as the substrate, which was re-added upon

116

deletion of both NH4+ and NO2-. All cultures were sub-cultured (10%, v/v) into autoclaved

117

fresh medium after three doses of the respective substrate (NH4Cl or NaNO2). Additional

118

pure AOB and AOA species (i.e., Nitrosomonas nitrosa Nm90 and Nitrososphaera

119

gargensis) were maintained as described previously.5

120

Biotransformation Experiments. The capability of MP biotransformation by the NOB

121

pure culture and the AOB/NOB co-culture were investigated in batch experiments.

122

Pre-grown biomass was harvested by centrifugation at 10,000 × g (4 °C for 15 min). The

123

biomass was re-suspended in autoclaved fresh medium and concentrated to reach a similar

124

ammonia or nitrite oxidation rate (~ 2 mM per 8 hours). Individual MP or mixed MP stock

125

solution was added to empty culture flasks. After the organic solvent methanol was

126

completely evaporated, 25 mL of thoroughly mixed concentrated NOB pure culture or

127

AOB/NOB co-culture was inoculated. As the mixture of tested MPs inhibited the

128

physiological activity of N. europaea at concentrations over 30 µg/L (each), and did not

129

show inhibition on Nitrobacter sp. at a concentration up to 100 µg/L (each), an initial MP

130

concentration of 100 µg/L (each) was used for the NOB experiments, while 10 µg/L (each)

131

was applied for the AOB/NOB co-culture. The culture flasks were shaken at 150 rpm for 5

132

min to re-dissolve the MPs. Weight loss due to evaporation in each flask was measured and 7 ACS Paragon Plus Environment


133

compensated by adding sterile nanopure water every 24 hours. Two mM NH4Cl and 2 mM

134

NaNO2 were added to the AOB/NOB co-culture and the NOB culture at the beginning,

135

respectively, and were re-added afterwards to keep the ammonia/nitrite oxidation activity.

136

For the AOB/NOB co-culture, 2 mM NH4Cl was re-added every 12 hours during the first

137

two days, and every 24 hours afterwards. For the NOB pure culture, 2 mM NaNO2 was

138

re-added every 24 hours. Samples (1 mL) were taken immediately after the first addition of

139

NH4Cl or NaNO2, and centrifuged at 16,000 × g (4 °C for 15 min). Supernatant was

140

collected and stored at 4 °C for LC-MS/MS and nitrogen species (i.e., NH4+, NO2-, and

141

NO3-) measurements. Cell pellets were stored at -20 °C for quantitative PCR (qPCR)

142

analysis. Samples were taken using the same procedure at subsequent time points during an

143

incubation period of 168 hours. At the end of incubation, cells in 10 mL culture were

144

collected by centrifugation, washed with sterile medium for three times, and stored at

145

-20 °C for intracellular MP and TP analysis. Biomass-free sterile fresh medium control and

146

heat-inactivated biomass (autoclavation at 121 °C for 20 min) control were set up in

147

triplicates in the same way as the biotransformation experiments described above. Briefly,

148

in both medium and heat-inactivated controls, the same MPs were added, and ammonium

149

(2 mM), nitrite (2 mM) and nitrate (14 mM) were added to mimic the levels of N-species

150

present in the biotransformation experiments.

151

Asulam Inhibitory Effect Tests. Biotransformation experiments were performed as

152

described above using the same pre-grown co-culture with different initial concentrations

153

of asulam (0, 10, 20, 30, 40, 50, and 80 µg/L). Two mM ammonia was supplied every 24

154

hours. Supernatant samples in a time series were taken over a 72-h incubation period, and 8 ACS Paragon Plus Environment

Page 8 of 39

Page 9 of 39


155

ammonia degradation and asulam biotransformation were determined as described above.

156

AMO Inhibition Experiments. To investigate the role of AMO in asulam

157

biotransformation, the bacterial AMO inhibitor octyne (~10 mg/L)8, 38 was added to the

158

AOB/NOB co-culture with 10 µg/L asulam after a 72-h incubation. We chose octyne as the

159

AMO inhibitor instead of the commonly used allylthiourea (ATU) and acetylene because (i)

160

ATU has been found not specific to AMO;8 (ii) acetylene is in gaseous phase, thus it is

161

difficult to keep its inhibitory effect in aerobic reactors with fast air exchange; (iii) octyne

162

is added as liquid, and was found to have more rapid inhibition than acetylene.38 All

163

inhibition experiments were performed in quadruplicates, three of which were sampled in a

164

time series over a 168-h incubation period for LC-MS/MS analysis, cell growth, and

165

N-species measurements. The remaining replicate was sampled every 24 hours to monitor

166

the ammonia oxidation activity using Hach test tubes for NH4- and NO3-N according to the

167

manufacturer’s instructions. Octyne was re-added as soon as recovery of ammonia

168

oxidation activity was detected.

169

Role of Ammonia Oxidation Intermediates in MP Transformation. To investigate the

170

effects of the ammonia oxidation intermediates, hydroxylamine (NH2OH) and nitric oxide

171

(NO) on MP biotransformation, MP transformation experiments after addition of either

172

NH2OH or NO were carried out in sterile fresh medium. Based on previously measured

173

concentrations of extracellular NH2OH produced by ammonia oxidizers in batch reactors

174

and the observation that 5 µM NH2OH was decomposed within 10 hours in the medium

175

used,39

176

comparison to AMO-mediated intermediate NH2OH, MP transformation by NO, a

5 µM NH2OH was added every 12 hours over a 168-h incubation period. As a

9 ACS Paragon Plus Environment


177

non-AMO-mediated intermediate was also carried out. Excess NO was added in the form

178

of 114.3 µM NONOate (Fisher Scientific, Hampton, NH), which releases ~ 200 µM NO

179

during the 168-h incubation (first-order process with half-life of 56 hours at 25 °C). A

180

mixture of all 17 MPs were added at a concentration of 20 µg/L for each compound.

181

Samples at 0, 12, and 168 h were taken for LC-HRMS analysis.

182

Analytical Method. MPs were analyzed by liquid chromatography coupled to a

183

high-resolution quadrupole orbitrap mass spectrometer (LC- HRMS/MS) (Q Exactive,

184

Thermo Fisher Scientific). For LC analysis, 50 µL sample was loaded onto a C18 Atlantis-

185

T3 column (particle size 3 µm, 3.0 × 150 mm, Waters), and eluted at a flow rate of 350

186

µL/min with nanopure water (A) and acetonitrile (B) (both amended with 0.1% formic

187

acid), at a gradient as follows: 5% B: 0 − 1 min, 5% − 100% B: 1 − 8min, 100% B: 8 − 20

188

min, and 5% B:20 − 26 min. The compounds were measured in full scan mode on HRMS

189

at a resolution of 70,000 at m/z 200 and a scan range of m/z 50 -750 in a positive/negative

190

switching mode. The limit of quantification (LOQ) for each MP is determined as the lowest

191

concentration with a detection variation < 20%, which was listed in Table S1.

192

Cell Extraction for Intracellular Concentration Measurement. A cell extraction

193

procedure from a previous study40 was used with slight modification. Briefly, internal

194

standards were spiked in cell pellets collected from 10 mL culture (at a final concentration

195

of 4 µg/L for each) followed by an addition of 2 mL lysis solvent containing methanol (0.5%

196

formic acid): nanopure water (0.1% w/w EDTA), 50: 50 (v/v). The cells were disrupted by

197

ultrasonication at 50 °C for 15 min, centrifuged at 10,000 × g for 10 min. The supernatant

198

was collected in a glass vial. This procedure was repeated twice for a better recovery. 10 ACS Paragon Plus Environment

Page 10 of 39

Page 11 of 39


199

Finally, ~ 6 mL supernatant was evaporated to dryness under a gentle steam of dinitrogen

200

gas at 40 °C. The analytes were re-dissolved in 0.5 mL filter-sterilized fresh medium

201

without CaCO3, which were then centrifuged at 10,000 × g at 4 °C for 10 min. The

202

supernatant was collected for LC-HRMS measurement.

203

Transformation Product Identification. Both suspect screening and non-target screening

204

were carried out to identify TPs. Suspect screening was done by TraceFinder 4.1 EFS

205

software (Thermo Fisher Scientific). TP suspect lists were compiled using an automated

206

metabolite mass prediction script, which considered a number of known redox and

207

hydrolysis reactions, as well as several conjugation reactions at primary and secondary

208

levels. Plausible TPs were identified according to the following criteria: i) mass accuracy

209

tolerance < 5ppm; ii) isotopic pattern score > 70%; iii) peak area > 5 × 106; iv) increasing

210

trend over time, or first increase then followed by a decrease; v) absent in biological

211

samples without MP addition or heat-inactivated controls. Sieve 2.2 software (Thermos

212

Fisher Scientific) was used for non-target screening and the TP candidates were selected

213

based on the same criteria. MS2 fragment profiles of TP candidates were obtained using

214

data-dependent MS/MS scan, to help elucidate TP structures. MarvinSketch (NET6.2.0,

215

2014) was used for drawing, displaying, and characterizing chemical structures,

216

ChemAxon (http://www.chemaxon.com).

217

Ammonium, Nitrite, and Nitrate Measurements. Ammonia/ammonium was measured

218

by a colorimetric method.41 Nitrite was measured by photometry with the sulfanilamide

219

N-(1-naphthyl) ethylenediamine dihydrochloride (NED) reagent method.42 Nitrate was first

220

reduced to nitrite by vanadium chloride, and then detected by the acidic Griess assay.43 The 11 ACS Paragon Plus Environment


221

detection limit for ammonium, nitrite, and nitrate was 10 µM.

222

Quantitative PCR. Genomic DNA was extracted by DNeasy Blood & Tissue Kit

223

(QIAGEN, Germantown, MD) according to manufacturer’s instructions. Cell growth was

224

measured by qPCR. Specific primers targeting the 16S rRNA genes of N. europaea,

225

(forward: 5’-GCAGCAGAGGGGAGTGGAAT-3’, reverse:

226

5’-CGTGCATGAGCGTCAGTGTC-3’), and Nitrobacter sp. (forward:

227

5’-CCATGACCGGTCGCAGAGAT-3’, reverse: 5’-AACTAAGGACGGGGGTTGCG-3’)

228

were designed and validated using the web-based tool, Primer-BLAST

229

(https://www.ncbi.nlm.nih.gov/tools/primer-blast/). FAST SYBR Green reagents (Thermo

230

Fisher Scientific) were used for qPCR according to the manufacturer’s instructions. In

231

general, each 20-µL reaction mixture contained 2.5 µL of gDNA sample or serially diluted

232

standard, 10 µL 2× Fast SYBR Green master mix, and 0.625 µM of the forward and reverse

233

primers. The PCR procedure included an initial deactivation at 95 °C for 20s, followed by

234

40 thermal cycles at 95 °C for 20s, then at 60 °C for 30s. Genomic DNA of N. europaea

235

and Nitrobacter sp. were quantified by NanoDrop One (Thermo Fisher Scientific) and

236

served as qPCR standards.

237

RESULTS AND DISCUSSION

238

MP Biotransformation by the AOB/NOB Co-culture and the NOB Pure Culture.

239

According to a previous study,8 the biotransformation of four selected compounds (i.e.,

240

asulam, monuron, clomazone, and trimethoprim) in a nitrifying activated sludge

241

community was highly associated with nitrification (50 − 90% inhibition when nitrification

242

was inhibited). To further demonstrate the causal relationship between nitrifiers and MP 12 ACS Paragon Plus Environment

Page 12 of 39

Page 13 of 39


243

biotransformation, we used an AOB/NOB co-culture that mimics the complete nitrification

244

system in WWTPs to investigate the biotransformation of the four selected MPs. This

245

co-culture setup also reduces the potential for biologically mediated abiotic nitritation that

246

has been observed for some MPs previously.33, 34 As AOB generally represent the dominant

247

ammonia-oxidizers in municipal WWTPs,15 the AOB model organism N. europaea was

248

selected to construct the co-culture with the NOB model organism Nitrobacter sp..

249

Nitrobacter frequently occurs in nitrifying activated sludge.44, 45 However, it should be

250

noticed that NOB of the genera Nitrotoga 44, 46, 47 and Nitrospira 48-50 often outnumber

251

Nitrobacter in nitrifying activated sludge systems. As biomass production for

252

biotransformation experiments is relatively slower with members of these genera, a

253

Nitrobacter strain was selected for the experiments, but future studies should also include

254

other NOB species. As expected, the co-culture converted all added ammonium to nitrate at

255

a conversion rate of 2 mM per 8 hours with no nitrite accumulation. A relative removal of

256

20% of a MP was arbitrarily selected as a threshold for a positive biotransformation result,

257

in order to make sure that other factors like measurement variations did not cause false

258

positive results. Among the four MPs, only asulam (as well as mianserin and ranitidine that

259

were included as positive controls for MP biotransformation) were biotransformed (> 20%

260

removal) by the co-culture during a 168-h incubation (Fig. 1A). Inconsistent with the

261

activated sludge inhibition study,8 no significant transformation of monuron, clomazone,

262

and trimethoprim was observed by the co-culture (Fig. S1). Intracellular asulam

263

concentrations of the co-culture in the biotransformation experiments were below the

264

detection limit. This excludes active biosorption/uptake by living cells, a phenomenon 13 ACS Paragon Plus Environment


265 266

observed for heavy metals.51 To elucidate whether the NOB species Nitrobacter sp. contributed to the

267

biotransformation in the co-culture, we tested the biotransformation of asulam, as well as

268

all the other 16 investigated MPs by the Nitrobacter sp. pure culture. Although the nitrite

269

oxidation of Nitrobacter sp. was not inhibited by the added MPs, there was no obvious

270

biotransformation (> 20% removal) for any of the tested MPs over a 168-h incubation

271

period (Fig. S2). Therefore, although NOB was co-inhibited with AOB in previous

272

inhibition studies due to the inhibited nitrite production,8, 10, 25, 52 they seem less likely

273

contribute to the biotransformation of the selected MPs, which showed a correlation with

274

nitrification activity. These results also indicate that the MP biotransformation in the

275

co-culture was carried out by the AOB species N. europaea.

276

Although biotransformation of monuron, clomazone, and trimethoprim was

277

strongly associated with nitrification in a previous study,8 these MPs were not

278

biotransformed by the ammonia oxidizers tested in this study. The inability of trimethoprim

279

biotransformation by N. europaea is consistent with findings by Khunjar et al.28 This

280

discrepancy between pure culture studies and community inhibition studies is possibly

281

because the nitrifying activated sludge in the inhibition study contained ammonia-oxidizers

282

with different biotransformation capabilities. This seems less likely, given the same

283

biotransformation compound specificity (i.e., of the four compounds, only asulam was

284

biotransformed at a 144-h removal of 25-40%) observed for the two AOB species (N.

285

europaea and Nitrosomonas nitrosa Nm90 isolated from activate sludge53) and the very

286

distantly related AOA species (Nitrososphaera gargensis)37 tested in this study (Fig. S3). 14 ACS Paragon Plus Environment

Page 14 of 39

Page 15 of 39


287

Nevertheless, another recently discovered ammonia oxidizer, comammox was also detected

288

in some WWTPs, which carries out complete ammonia oxidation to nitrate, and is

289

physiologically distinct from the other ammonia oxidizers (i.e., AOB and AOA).54

290

Although beyond the scope of this study, it is worthwhile to investigate their roles in MP

291

biotransformation in future studies. Another possible reason for the discrepancy between

292

pure culture studies and community inhibition studies is that NOB species other than

293

Nitrobacter, or heterotrophic microorganisms growing dependently on nitrifiers55 in the

294

nitrifying activated sludge were responsible for the biotransformation of the MPs that were

295

not transformed by the tested ammonia oxidizers. For example, studies have indicated that

296

trimethoprim can undergo various biotransformation pathways under both aerobic

297

(nitrifying) and anoxic (denitrifying) conditions.56 Therefore, it is highly likely that

298

heterotrophic microorganisms physiologically interacting with nitrifiers carried out the

299

biotransformation of trimethoprim in nitrifying activated sludge communities. Although

300

lack of previous bacterial biotransformation studies, the biotransformation of monuron and

301

clomazone, similar with trimethoprim biotransformation was likely carried out by

302

heterotrophs.

303

Transformation Product Identification. To identify TPs of asulam, suspect screening was

304

first carried out. The self-compiled suspect list includes ~ 1000 compounds derived from

305

asulam via various reactions, including conjugation at primary and secondary levels. By

306

suspect screening, two asulam TPs were identified in the nitrifying co-culture, with exact

307

masses of [M-H] at 230.0129 (denoted “TP230”), and at 274.9979 (denoted “TP274”) (Fig.

308

1B). The formula of TP230 is C8H9O5NS (−NH+O from asulam), and the formula of 15 ACS Paragon Plus Environment


309

TP274 is C8H8O7N2S (−2H+3O from asulam). Both TPs were likely formed from oxidation

310

reactions, and their hypothetical structures are proposed (Fig. S4). TP230 was probably

311

formed by the oxidation of the amine group (−NH2) of asulam to a hydroxyl group (−OH),

312

and TP274 likely contains a nitro group from oxidation of the amine group, as well as a

313

hydroxyl group added on the aromatic ring (Fig. S4). It is worth noting that the removal of

314

asulam did not lead to a complete breakdown of the main structure into simpler molecules,

315

as both identified TPs have similar or even larger masses than asulam. MS2 analysis were

316

further carried out to validate the proposed structures of the two TPs (Fig. S4). However,

317

the MS2 fragments of the two TPs did not provide clear evidence for the position of the

318

adding −OH or nitro group. As the corresponding standard compounds with the proposed

319

TP structures were unavailable, there were no reference MS2 fragmentation profiles to

320

compare with, thus the exact TP structures remain unclear.

321

According to the peak areas, the two TPs account for ~ 40% of the removed asulam,

322

resulting in an incomplete mass balance. Therefore, we further carried out non-target

323

screening to identify TP candidates that might be missing from the suspect screening.

324

However, except the above two TPs, no additional plausible TP was detected. We also

325

analyzed possible TPs in the intracellular samples using suspect and non-target screening.

326

According to the same screening criteria, no TP candidate was detected in the intracellular

327

samples, which made it unlikely of active uptake of TPs into living cells. As asulam is a

328

relatively small molecule (m/z 231.0434, [M+H]), possible reasons for the incomplete mass

329

balance include: (1) formation of even smaller molecules that are not detectable by LC-MS;

330

(2) asulam and the two TPs were detected in two ESI modes (positive for asulam, and 16 ACS Paragon Plus Environment

Page 16 of 39

Page 17 of 39


331

negative for the TPs). The ionization efficiency under negative mode is usually lower than

332

positive mode, resulting in lower response intensities for the same concentration of

333

compounds.

334

The incomplete biotransformation of asulam and poor mass balance nicely illustrate

335

that it is crucial to elucidate TP profiles during transformation processes to obtain a

336

comprehensive understanding of the environmental fate of parent MPs. Future research is

337

needed to better understand the ecological impact and putative health risks associated with

338

TPs from incomplete MP transformation.

339

Co-metabolic Biotransformation of Asulam. To investigate the biotransformation

340

mechanism of asulam by N. europaea in the nitrifying co-culture, limited ammonium

341

(one-time addition of 0.05 mM NH4Cl at the beginning of the experiment) was added to the

342

co-culture only as the nitrogen source, and asulam was added as a putative energy source.

343

Under these conditions, no asulam biotransformation was observed (Fig. 2A), with no

344

ammonia oxidation (Fig. 2B) and little cell growth (Fig. 2C). In comparison, asulam was

345

continuously biotransformed with active ammonium oxidation and cell growth in the

346

control with ammonium as the energy and nitrogen source (Fig. 2). These results suggest

347

that asulam biotransformation by N. europaea in the co-culture occurred via co-metabolism.

348

Similarly, previous studies have shown that AOB can co-metabolically biotransform a

349

number of other MPs, which has often been hypothesized to be carried out by AMO given

350

its broad substrate spectrum.5, 11, 12, 32, 57-59 However, experimental support for this

351

hypothesis is scarce.

352

Experimental Support for AMO as the Responsible Enzyme for Asulam 17 ACS Paragon Plus Environment


353

Biotransformation. To investigate whether AMO is the responsible enzyme for asulam

354

biotransformation, inhibitory effects of asulam on ammonia oxidation was first examined.

355

Compared to the no-asulam control, no inhibition of ammonia oxidation was observed for

356

asulam at 10, 20, and 30 µg/L (Fig. S5). The ammonia oxidation rate started to decrease

357

with 40 µg/L asulam by 4.7%, and the inhibition increased up to 15.5% at a concentration

358

of 80 µg/L (Fig. S5). Interestingly, the inhibitory effect of asulam on ammonia oxidation

359

decreased gradually after re-feeding the co-culture with ammonium (Fig. S5). This is

360

possibly because, the ratio of ammonia to asulam increased after re-feeding with

361

ammonium as asulam was removed over time which led to a decrease in potential substrate

362

competition for AMO. Moreover, newly synthesized AMO during cell growth could also

363

lead to the recovery of ammonia oxidation with a longer incubation. The inhibitory effect

364

of asulam on ammonia oxidation suggested that asulam might be a competitive substrate

365

for AMO, which was likely the responsible enzyme for asulam biotransformation. A

366

substantial inhibitory effect of asulam was observed at 80 µg/L (c.a., 348 nM) (Fig. S5),

367

indicating that asulam possibly has an even higher affinity to AMO than ammonia (the

368

ammonia affinity of the AMO of N. europaea is reported between 20 − 70 µM NH3, c.a., ~

369

330 − 1200 µM of total ammonium at pH 8).54

370

In addition, a specific mechanistic inhibitor of bacterial AMO, octyne,35, 36 was used

371

to further examine whether AMO was the responsible enzyme for the cometabolic asulam

372

biotransformation. Octyne was added after 72 hours during asulam biotransformation in the

373

nitrifying co-culture. As the addition of octyne resulted in instantaneous inhibition of AMO

374

activity, we hypothesized that, if AMO was the responsible enzyme, immediate inhibition 18 ACS Paragon Plus Environment

Page 18 of 39

Page 19 of 39


375

of asulam biotransformation would occur after octyne addition, while if other enzymes

376

carried out the biotransformation, a lagged inhibition would be expected as their activities

377

likely persist for a while after AMO inhibition.30 Ammonia oxidation and asulam

378

transformation was determined 2 hours after the octyne addition. Compared to the

379

no-octyne-treated control, in the octyne-treated co-culture, complete inhibition of ammonia

380

oxidation was observed (Fig. 3A), and asulam biotransformation also completely ceased

381

(Fig. 3B). Consistently, no growth of AOB and NOB occurred after octyne exposure (Fig.

382

3C&D). These results provide strong support for the assumption that AMO is the enzyme

383

involved in asulam biotransformation by N. europaea in the co-culture.

384

Abiotic MP Transformation by NH2OH, the Intermediate Formed by AMO. In

385

addition to being converted by AMO as non-specific substrates, it is also possible that MPs

386

abiotically react with hydroxylamine, the product of ammonia oxidation by AMO.39 In

387

order to test this, we conducted abiotic MP transformation experiments with the addition of

388

NH2OH at concentrations occurring in AOB batch cultures.39 We also tested MP

389

transformation by NO, a non-AMO-mediated ammonia oxidation intermediate of AOB.

390

NO can be formed by nitrite reductase (NirK) in AOB.60-62 Moreover, it is produced as an

391

obligate intermediate by hydroxylamine dehydrogenase (HAO) of AOB. 63, 64 All 15 MPs

392

and the two control MPs mianserin and ranitidine were exposed to hydroxylamine and NO,

393

respectively. The abiotic transformation was compared to the biological transformation by

394

the AOB/NOB co-culture (Fig. 4). Eight MPs were biotransformed by the AOB/NOB

395

co-culture (Fig. 4, Fig. S6), whereas no biotransformation was observed in the tested NOB

396

pure culture for all MPs (Fig. S2). Six out of the eight MPs were also transformed 19 ACS Paragon Plus Environment


397

abiotically in NH2OH-added fresh medium (with even higher removals), and only

398

rufinamide and bezafibrate did not exhibit transformation by NH2OH (Fig. 4). In contrast,

399

only three of the eight MPs, i.e., indomethacin, ranitidine and fenhexamid, were

400

transformed by NO, taken into consideration the abiotic removal of mianserin and

401

indomethacin (61% and 22%, respectively) in fresh medium with no NH2OH or NO

402

addition (Fig. 4). Transformation of those three compounds by NO were less pronounced

403

than by NH2OH (Fig. 4). Furthermore, one cannot exclude the possibility that the

404

transformation of the three MPs was attributed to reactions with the NONOate residual

405

rather than the released NO.

406

Further analyses on TPs focused on MPs transformed by NH2OH. For asulam,

407

TP230 and TP274 identified during biotransformation in the co-culture were also detected

408

in the NH2OH-treated fresh medium. The similar compound specificities and TP profiles

409

between biotransformation and NH2OH-mediated abiotic transformation suggest that

410

abiotic MP transformation by NH2OH might be an important transformation route

411

indirectly mediated by AMO, which was not recognized previously. The abundance of

412

TP230 in the NH2OH-treated fresh medium after 168 hours was higher than that in the

413

co-culture when adding the same concentration of asulam (20 µg/L) (Fig. 5). One should

414

notice that more asulam was converted in the NH2OH-treated fresh medium (> 80%) than

415

in the co-culture (~ 60%) after 168 hours (Fig. 4), which might result in the higher

416

formation of TP230 in the NH2OH-treated fresh medium. In other words, if asulam

417

removal in the co-culture reached the same level as that in the NH2OH-treated fresh

418

medium, the abundance of TP230 under both conditions would be similar. Under the 20 ACS Paragon Plus Environment

Page 20 of 39

Page 21 of 39


419

assumption that the abundance of TPs formed from non-enzymatic transformation by

420

NH2OH in the co-culture was similar as that observed in fresh medium with NH2OH

421

addition, this result suggests that the formation of TP230 was more likely via the

422

NH2OH-mediated abiotic transformation route. Nevertheless, the abundance of TP274 in

423

the AOB/NOB co-culture was substantially higher than that in the NH2OH-treated fresh

424

medium (Fig. 5), which indicates the co-occurrence of non-enzymatic and enzymatic MP

425

transformation pathways for the formation of TP274.

426

Among the other five MPs (i.e. mianserin, indomethacin, ranitidine, furosemide,

427

and fenhexamid) transformed abiotically in NH2OH-added fresh medium, no plausible TP

428

could be detected for indomethacin, furosemide, and fenhexamid in both the AOB/NOB

429

co-culture and the NH2OH-added fresh medium. For mianserin and ranitidine, consistent

430

with the previous study,5 the same TPs: TP279 for mianserin, and TP273, TP289, TP330

431

for ranitidine were detected in the AOB/NOB co-culture, as well as in the NH2OH-added

432

fresh medium, but at lower concentrations compared to those in the AOB/NOB co-culture

433

(Fig. S7). One possible reason is that in the AOB/NOB co-culture the same TPs were

434

formed via both AMO-mediated biotransformation and NH2OH-mediated abiotic

435

transformation, leading to a higher TP formation.

436

Moreover, by suspect screening, two major TPs of rufinamide with exact masses of

437

[M + H] at 240.0597 (TP240), and 312.0790 (TP312) (Fig. S8) detected in the AOB/NOB

438

co-culture were not detected in the NH2OH-treated fresh medium, which was in agreement

439

with no transformation of rufinamide by NH2OH abiotically (Fig. 4).

440

With NH2OH treatment, it is surprising that the identified TPs were in oxidized 21 ACS Paragon Plus Environment


441

forms of the respective MPs, given that NH2OH is nucleophilic and typically serves as a

442

reducing agent. Moreover, no reduced forms of asulam TPs could be detected when

443

targeting the corresponding TP suspects based on reported reductive reactions of NH2OH

444

with compounds that also have carboxylic ester functional groups as asulam.65 The

445

oxidized forms of TPs were probably formed from reactions of the MPs with some

446

oxidizing intermediates from the reaction of NH2OH with dissolved O2. In this context, it is

447

interesting to note that although the initial dose of NH2OH (5 µM) was in excess compared

448

to the MP concentrations (~ 0.06 - 0.12 µM), the abiotic transformation of MPs by NH2OH

449

did not reach the highest level after 12 hours with the addition of a single dose of NH2OH,

450

but further increased upon reamendment of NH2OH during a longer exposure (Fig. 4). This

451

could be due to faster degradation of NH2OH when reacting with dissolved O2 than reacting

452

with the MPs, or due to the possibility that the MPs reacted with intermediates from

453

NH2OH decomposition in water, which could be formed at lower concentrations. However,

454

the exact pathways and mechanisms of NH2OH-mediated abiotic MP transformation need

455

to be further elucidated.

456

Environmental Relevance and Implications. Nitrifiers are ubiquitously distributed in

457

natural environments, as well as engineered systems such as WWTPs. Although positive

458

correlations have been previously observed between MP biotransformation and nitrification

459

activities,8 the responsible enzymes and involved pathways are not well understood. Here,

460

we demonstrate that the NOB Nitrobacter sp. did not contribute to the biotransformation of

461

all tested MPs whose biotransformation was associated with nitrification activity according

462

to nitrifying activated sludge inhibition studies.8 We also show that only 6 of 15 MPs, 22 ACS Paragon Plus Environment

Page 22 of 39

Page 23 of 39


463

whose biotransformation positively correlated with the activity of nitrifiers in a nitrifying

464

activated sludge community were transformed by the AOB species in an AOB/NOB (N.

465

europaea/Nitrobacter sp.) co-culture. Similar results were observed in other studies,11, 28, 52

466

where more MPs were biotransformed by complex communities such as nitrifying

467

activated sludge than by various pure ammonia-oxidizing cultures, including AOB and

468

AOA. Besides potential contributions of other physiologically diverse ammonia oxidizers

469

and NOB, including the recently discovered complete ammonia oxidizers20, 21 to MP

470

biotransformation, it is highly likely that heterotrophic bacteria interactively co-existing

471

with nitrifiers also make important contributions to MP transformation in nitrifying

472

activated sludge. Thus, keeping the diversity and complexity in biological systems could be

473

important to obtain a higher overall removal of MPs. Further studies are needed to

474

investigate effects of microbial diversity on MP biotransformation, and to identify the key

475

heterotrophs that co-exist with nitrifiers and contribute to MP biotransformation.

476

Importantly, we present evidence favoring that AMO of N. europaea is the

477

responsible enzyme for asulam biotransformation, which likely occurred via co-metabolism.

478

Furthermore, we report for the first time that abiotic MP transformation by hydroxylamine,

479

the product of AMO-mediated ammonia oxidation, likely represents another transformation

480

mechanism of MPs by ammonia oxidizers. It expands our knowledge on MP

481

transformation pathways and highlights that reactive intermediates of different

482

environmental microbial guilds might play an important yet understudied role in MP

483

transformation. As an NH4-N concentration (2 mM, c.a. 28 mg/L as N) relevant to typical

484

influent of municipal WWTPs was used in the batch reactors of this study, similar MP 23 ACS Paragon Plus Environment


485

transformation pathways might be occurring in nitrifying activated sludge in continuous

486

flow reactors in WWTPs. However, it needs to be further investigated whether the actual

487

free NH2OH available for abiotic MP transformation in WWTPs was similar with that in

488

lab-scale batch reactors. Altogether, these findings give a better understanding of roles

489

played by nitrifiers (i.e., AOB and NOB) in MP biotransformation, the responsible enzyme,

490

and the underlying mechanisms, which will provide important insights into the fate of MPs

491

in various nitrifying environments.

492

SUPPORTING INFORMATION

493

The Supporting Information is available on the ACS Publication Website, including Table

494

S1, and Fig. S1-S8.

495

ACKNOWLEDGEMENTS

496

We would like to thank Mengwei Han from the Department of Civil and Environmental

497

Engineering at University of Illinois at Urbana-Champaign and Dr. Zhenyu Tian from the

498

Center for Urban Waters at University of Washington Tacoma for their thoughtful

499

discussion on LC-MS/MS data analysis. P.H. and M.W. were supported by the European

500

Research Council Advanced Grant project NITRICARE 294343. L.J.Z. was supported by

501

the National Natural Science Foundation of China (31770551).

502

REFERENCES

503

(1) Schwarzenbach, R. P.; Escher, B. I.; Fenner, K.; Hofstetter, T. B.; Johnson, C. A.; von

504

Gunten, U.; Wehrli, B. The challenge of micropollutants in aquatic systems. Science 2006,

505

313, 1072-1077.

506

(2) Bolong, N.; Ismail, A. F.; Salim, M. R.; Matsuura, T. A review of the effects of 24 ACS Paragon Plus Environment

Page 24 of 39

Page 25 of 39


507

emerging contaminants in wastewater and options for their removal. Desalination 2009,

508

239, 229-246.

509

(3) Fenner, K.; Canonica, S.; Wackett, L. P.; Elsner, M. Evaluating pesticide degradation in

510

the environment: blind spots and emerging opportunities. Science 2013, 341, 752-758.

511

(4) Daughton, C. G.; Ternes, T. A. Pharmaceuticals and personal care products in the

512

environment: agents of subtle change? Environ. Health Perspect. 1999, 107 Suppl 6,

513

907-938.

514

(5) Men, Y. J.; Han, P.; Helbling, D. E.; Jehmlich, N.; Herbold, C.; Gulde, R.;

515

Onnis-Hayden, A.; Gu, A. Z.; Johnson, D. R.; Wagner, M.; Fenner, K. Biotransformation of

516

two pharmaceuticals by the ammonia-oxidizing archaeon Nitrososphaera gargensis.

517

Environ. Sci. Technol. 2016, 50, 4682-4692.

518

(6) Nguyen, L. N.; Hai, F. I.; Dosseto, A.; Richardson, C.; Price, W. E.; Nghiem, L. D.

519

Continuous adsorption and biotransformation of micropollutants by granular activated

520

carbon-bound laccase in a packed-bed enzyme reactor. Bioresour. Technol. 2016, 210,

521

108-116.

522

(7) Luo, Y. L.; Guo, W. S.; Ngo, H. H.; Nghiem, L. D.; Hai, F. I.; Zhang, J.; Liang, S.;

523

Wang, X. C. C. A review on the occurrence of micropollutants in the aquatic environment

524

and their fate and removal during wastewater treatment. Sci. Total Environ. 2014, 473-474,

525

619-41.

526

(8) Men, Y. J.; Achermann, S.; Helbling, D. E.; Johnson, D. R.; Fenner, K. Relative

527

contribution of ammonia oxidizing bacteria and other members of nitrifying activated

528

sludge communities to micropollutant biotransformation. Water Res. 2017, 109, 217-226. 25 ACS Paragon Plus Environment


529

(9) Clara, M.; Kreuzinger, N.; Strenn, B.; Gans, O.; Kroiss, H. The solids retention time - a

530

suitable design parameter to evaluate the capacity of wastewater treatment plants to remove

531

micropollutants. Water Res. 2005, 39, 97-106.

532

(10) Fernandez-Fontaina, E.; Gomes, I. B.; Aga, D. S.; Omil, F.; Lema, J. M.; Carballa, M.

533

Biotransformation of pharmaceuticals under nitrification, nitratation and heterotrophic

534

conditions. Sci. Total Environ. 2016, 541, 1439-1447.

535

(11) Xu, Y. F.; Yuan, Z. G.; Ni, B. J. Biotransformation of pharmaceuticals by ammonia

536

oxidizing bacteria in wastewater treatment processes. Sci. Total Environ. 2016, 566,

537

796-805.

538

(12) Margot, J.; Lochmatter, S.; Barry, D. A.; Holliger, C. Role of ammonia-oxidizing

539

bacteria in micropollutant removal from wastewater with aerobic granular sludge. Water

540

Sci. Technol. 2016, 73, 564-575.

541

(13) Stamatis, N. K.; Konstantinou, I. K. Occurrence and removal of emerging

542

pharmaceutical, personal care compounds and caffeine tracer in municipal sewage

543

treatment plant in Western Greece. J. Environ. Sci. Health, Part B 2013, 48, 800-813.

544

(14) Estevez, E.; Cabrera, M. D.; Molina-Diaz, A.; Robles-Molina, J.; Palacios-Diaz, M. D.

545

Screening of emerging contaminants and priority substances (2008/105/EC) in reclaimed

546

water for irrigation and groundwater in a volcanic aquifer (Gran Canaria, Canary Islands,

547

Spain). Sci. Total Environ. 2012, 433, 538-546.

548

(15) Mußmann, M.; Brito, I.; Pitcher, A.; Sinninghe Damsté, J. S.; Hatzenpichler, R.;

549

Richter, A.; Nielsen, J. L.; Nielsen, P. H.; Müller, A.; Daims, H.; Wagner, M.; Head, I. M.

550

Thaumarchaeotes abundant in refinery nitrifying sludges express amoA but are not obligate 26 ACS Paragon Plus Environment

Page 26 of 39

Page 27 of 39


551

autotrophic ammonia oxidizers. Proc. Natl. Acad. Sci. U. S. A. 2011, 108, 16771-16776.

552

(16) Li, Y.; Ding, K.; Wen, X.; Zhang, B.; Shen, B.; Yang, Y. A novel ammonia-oxidizing

553

archaeon from wastewater treatment plant: its enrichment, physiological and genomic

554

characteristics. Sci. Rep. 2016, 6, 23747.

555

(17) Limpiyakorn, T.; Sonthiphand, P.; Rongsayamanont, C.; Polprasert, C. Abundance of

556

amoA genes of ammonia-oxidizing archaea and bacteria in activated sludge of full-scale

557

wastewater treatment plants. Bioresour. Technol. 2011, 102, 3694-3701.

558

(18) Park, H. D.; Wells, G. F.; Bae, H.; Criddle, C. S.; Francis, C. A. Occurrence of

559

ammonia-oxidizing archaea in wastewater treatment plant bioreactors. Appl. Environ.

560

Microbiol. 2006, 72, 5643-5647.

561

(19) Pjevac, P.; Schauberger, C.; Poghosyan, L.; Herbold, C. W.; van Kessel, M. A. H. J.;

562

Daebeler, A.; Steinberger, M.; Jetten, M. S. M.; Lücker, S.; Wagner, M.; Daims, H.

563

AmoA-targeted polymerase chain reaction primers for the specific detection and

564

quantification of comammox Nitrospira in the environment. Front. Microbiol. 2017, 8,

565

1508.

566

(20) Daims, H.; Lebedeva, E. V.; Pjevac, P.; Han, P.; Herbold, C.; Albertsen, M.; Jehmlich,

567

N.; Palatinszky, M.; Vierheilig, J.; Bulaev, A.; Kirkegaard, R. H.; von Bergen, M.; Rattei, T.;

568

Bendinger, B.; Nielsen, P. H.; Wagner, M. Complete nitrification by Nitrospira bacteria.

569

Nature 2015, 528, 504-509.

570

(21)van Kessel, M. A. H. J.; Speth, D. R.; Albertsen, M.; Nielsen, P. H.; Op den Camp, H. J.

571

M.; Kartal, B.; Jetten, M. S. M.; Lücker, S. Complete nitrification by a single

572

microorganism. Nature 2015, 528, 555-559. 27 ACS Paragon Plus Environment


573

(22) Sathyamoorthy, S.; Chandran, K.; Ramsburg, C. A. Biodegradation and cometabolic

574

modeling of selected beta blockers during ammonia oxidation. Environ. Sci. Technol. 2013,

575

47, 12835-12843.

576

(23) Batt, A. L.; Kim, S.; Aga, D. S. Enhanced biodegradation of iopromide and

577

trimethoprim in nitrifying activated sludge. Environ. Sci. Technol. 2006, 40, 7367-7373.

578

(24) Gobel, A.; McArdell, C. S.; Joss, A.; Siegrist, H.; Giger, W. Fate of sulfonamides,

579

macrolides, and trimethoprim in different wastewater treatment technologies. Sci. Total

580

Environ. 2007, 372, 361-371.

581

(25) Tran, N. H.; Urase, T.; Kusakabe, O. The characteristics of enriched nitrifier culture in

582

the degradation of selected pharmaceutically active compounds. J. Hazard. Mater. 2009,

583

171, 1051-1057.

584

(26) Dorival-Garcia, N.; Zafra-Gomez, A.; Navalon, A.; Gonzalez-Lopez, J.; Hontoria, E.;

585

Vilchez, J. L. Removal and degradation characteristics of quinolone antibiotics in

586

laboratory-scale activated sludge reactors under aerobic, nitrifying and anoxic conditions. J.

587

Environ. Manage. 2013, 120, 75-83.

588

(27) Rattier, M.; Reungoat, J.; Keller, J.; Gernjak, W. Removal of micropollutants during

589

tertiary wastewater treatment by biofiltration: role of nitrifiers and removal mechanisms.

590

Water Res. 2014, 54, 89-99.

591

(28) Khunjar, W. O.; Mackintosh, S. A.; Skotnicka-Pitak, J.; Baik, S.; Aga, D. S.; Love, N.

592

G. Elucidating the relative roles of ammonia oxidizing and heterotrophic bacteria during

593

the biotransformation of 17 alpha-ethinylestradiol and trimethoprim. Environ. Sci. Technol.

594

2011, 45, 3605-3612. 28 ACS Paragon Plus Environment

Page 28 of 39

Page 29 of 39


595

(29) Yi, T.; Harper, W. F. The link between nitrification and biotransformation of

596

17α-ethinylestradiol. Environ. Sci. Technol. 2007, 41, 4311-4316.

597

(30) Arp, D. J.; Sayavedra-Soto, L. A.; Hommes, N. G. Molecular biology and biochemistry

598

of ammonia oxidation by Nitrosomonas europaea. Arch. Microbiol. 2002, 178, 250-255.

599

(31) Rasche, M. E.; Hicks, R. E.; Hyman, M. R.; Arp, D. J. Oxidation of monohalogenated

600

ethanes and n-chlorinated alkanes by whole cells of Nitrosomonas europaea. J. Bacteriol.

601

1990, 172, 5368-5373.

602

(32) Keener, W. K.; Arp, D. J. Transformations of aromatic compounds by Nitrosomonas

603

europaea. Appl. Environ. Microbiol. 1994, 60, 1914-1920.

604

(33) Sun, Q.; Li, Y.; Chou, P. H.; Peng, P. Y.; Yu, C. P. Transformation of bisphenol A and

605

alkylphenols by ammonia-oxidizing bacteria through nitration. Environ. Sci. Technol. 2012,

606

46, 4442-4448.

607

(34) Gaulke, L. S.; Strand, S. E.; Kalhorn, T. F.; Stensel, H. D. 17 α-ethinylestradiol

608

transformation via abiotic nitration in the presence of ammonia oxidizing bacteria. Environ.

609

Sci. Technol. 2008, 42, 7622-7627.

610

(35) Taylor, A. E.; Taylor, K.; Tennigkeit, B.; Palatinszky, M.; Stieglmeier, M.; Myrold, D.

611

D.; Schleper, C.; Wagner, M.; Bottomley, P. J. Inhibitory effects of C-2 to C-10 1-alkynes

612

on ammonia oxidation in two Nitrososphaera species. Appl. Environ. Microbiol. 2015, 81,

613

1942-1948.

614

(36) Taylor, A. E.; Vajrala, N.; Giguere, A. T.; Gitelman, A. I.; Arp, D. J.; Myrold, D. D.;

615

Sayavedra-Soto, L.; Bottomley, P. J. Use of aliphatic n-alkynes to discriminate soil

616

nitrification activities of ammonia-oxidizing thaumarchaea and bacteria. Appl. Environ. 29 ACS Paragon Plus Environment


617

Microbiol. 2013, 79, 6544-6551.

618

(37) Palatinszky, M.; Herbold, C.; Jehmlich, N.; Pogoda, M.; Han, P.; von Bergen, M.;

619

Lagkouvardos, I.; Karst, S. M.; Galushko, A.; Koch, H.; Berry, D.; Daims, H.; Wagner, M.

620

Cyanate as an energy source for nitrifiers. Nature 2015, 524, 105-108.

621

(38) Hyman, M. R.; Murton, I. B.; Arp, D. J. Interaction of ammonia monooxygenase from

622

Nitrosomonas europaea with alkanes, alkenes, and alkynes. Appl. Environ. Microbiol. 1988,

623

54, 3187-3190.

624

(39) Liu, S.; Han, P.; Hink, L.; Prosser, J. I.; Wagner, M.; Brüggemann, N. Abiotic

625

conversion of extracellular NH2OH contributes to N2O emission during ammonia oxidation.

626

Environ. Sci. Technol. 2017, 22, 13122-13132.

627

(40) Gago-Ferrero, P.; Borova, V.; Dasenaki, M. E.; Thomaidis, N. S. Simultaneous

628

determination of 148 pharmaceuticals and illicit drugs in sewage sludge based on

629

ultrasound-assisted extraction and liquid chromatography-tandem mass spectrometry. Anal.

630

Bioanal. Chem. 2015, 407, 4287-4297.

631

(41) Kandeler, E.; Gerber, H. Short-term assay of soil urease activity using colorimetric

632

determination of ammonium. Biol. Fertil. Soils 1988, 6, 68-72.

633

(42) Strickland, J. D. H.; Parsons, T. R. A practical handbook of seawater analysis. Bulletin

634

167: Fisheries Research Board of Canada: Ottawa, 1972.

635

(43) Miranda, K. M.; Espey, M. G.; Wink, D. A. A rapid, simple spectrophotometric method

636

for simultaneous detection of nitrate and nitrite. Nitric Oxide Biol. Ch. 2001, 5, 62-71.

637

(44) Alawi, M.; Off, S.; Kaya, M.; Spieck, E. Temperature influences the population

638

structure of nitrite-oxidizing bacteria in activated sludge. Environ. Microbiol. Rep. 2009, 1, 30 ACS Paragon Plus Environment

Page 30 of 39

Page 31 of 39


639

184-190.

640

(45) Mobarry, B. K.; Wagner, M.; Urbain, V.; Rittmann, B. E.; Stahl, D. A. Phylogenetic

641

probes for analyzing abundance and spatial organization of nitrifying bacteria. Appl.

642

Environ. Microbiol. 1996, 62, 2156-2162.

643

(46) Saunders, A. M.; Albertsen, M.; Vollertsen, J.; Nielsen, P. H. The activated sludge

644

ecosystem contains a core community of abundant organisms. ISME J. 2016, 10, 11-20.

645

(47) Lücker, S.; Schwarz, J.; Gruber-Dorninger, C.; Spieck, E.; Wagner, M.; Daims, H.

646

Nitrotoga-like bacteria are previously unrecognized key nitrite oxidizers in full-scale

647

wastewater treatment plants. ISME J. 2015, 9, 708-720.

648

(48) Gruber-Dorninger, C.; Pester, M.; Kitzinger, K.; Savio, D. F.; Loy, A.; Rattei, T.;

649

Wagner, M.; Daims, H. Functionally relevant diversity of closely related Nitrospira in

650

activated sludge. ISME J. 2015, 9, 643-655.

651

(49) Daims, H.; Nielsen, J. L.; Nielsen, P. H.; Schleifer, K. H.; Wagner, M. In situ

652

characterization of Nitrospira-like nitrite-oxidizing bacteria active in wastewater treatment

653

plants. Appl. Environ. Microbiol. 2001, 67, 5273-5284.

654

(50) Juretschko, S.; Timmermann, G.; Schmid, M.; Schleifer, K. H.; Pommerening-Röser,

655

A.; Koops, H. P.; Wagner, M. Combined molecular and conventional analyses of nitrifying

656

bacterium diversity in activated sludge: Nitrosococcus mobilis and Nitrospira-like bacteria

657

as dominant populations. Appl. Environ. Microbiol. 1998, 64, 3042-3051.

658

(51) You, S. J.; Tsai, Y. P.; Huang, R. Y. Effect of heavy metals on nitrification performance

659

in different activated sludge processes. J. Hazard. Mater. 2009, 165, 987-994.

660

(52) Roh, H.; Subramanya, N.; Zhao, F.; Yu, C. P.; Sandt, J.; Chu, K. H. Biodegradation 31 ACS Paragon Plus Environment


661

potential of wastewater micropollutants by ammonia-oxidizing bacteria. Chemosphere

662

2009, 77, 1084-1089.

663

(53) Koops, H. P.; Böttcher, B.; Möller, U. C.; Pommerening-Röser, A.; Stehr, G.

664

Classification of eight new species of ammonia-oxidizing bacteria: Nitrosomonas

665

communis sp. nov., Nitrosomonas ureae sp. nov., Nitrosomonas aestuarii sp. nov.,

666

Nitrosomonas marina sp. nov., Nitrosomonas nitrosa sp. nov., Nitrosomonas eutropha sp.

667

nov., Nitrosomonas oligotropha sp. nov. and Nitrosomonas halophila sp. nov. J. Gen.

668

Microbial. 1991, 137, 1689-1699.

669

(54) Kits, K. D.; Sedlacek, C. J.; Lebedeva, E. V.; Han, P.; Bulaev, A.; Pjevac, P.; Daebeler,

670

A.; Romano, S.; Albertsen, M.; Stein, L. Y.; Daims, H.; Wagner, M. Kinetic analysis of a

671

complete nitrifier reveals an oligotrophic lifestyle. Nature 2017, 549, 269-272.

672

(55) Dolinsek, J.; Lagkouvardos, I.; Wanek, W.; Wagner, M.; Daims, H. Interactions of

673

nitrifying bacteria and heterotrophs: identification of a Micavibrio-like putative predator of

674

Nitrospira spp. Appl. Environ. Microbiol. 2013, 79, 2027-2037.

675

(56) Jewell, K. S.; Castronovo, S.; Wick, A.; Falas, P.; Joss, A.; Ternes, T. A. New insights

676

into the transformation of trimethoprim during biological wastewater treatment. Water Res.

677

2016, 88, 550-557.

678

(57) Xu, Y. F.; Radjenovic, J.; Yuan, Z. G.; Ni, B. J. Biodegradation of atenolol by an

679

enriched nitrifying sludge: products and pathways. Chem. Eng. J. 2017, 312, 351-359.

680

(58) Quintana, J. B.; Weiss, S.; Reemtsma, T. Pathways and metabolites of microbial

681

degradation of selected acidic pharmaceutical and their occurrence in municipal wastewater

682

treated by a membrane bioreactor. Water Res. 2005, 39, 2654-2664. 32 ACS Paragon Plus Environment

Page 32 of 39

Page 33 of 39


683

(59) Tran, N. H.; Urase, T.; Ngo, H. H.; Hu, J.; Ong, S. L. Insight into metabolic and

684

cometabolic activities of autotrophic and heterotrophic microorganisms in the

685

biodegradation of emerging trace organic contaminants. Bioresour. Technol. 2013, 146,

686

721-731.

687

(60) Kozlowski, J. A.; Price, J.; Stein, L. Y. Revision of N2O-producing pathways in the

688

ammonia-oxidizing bacterium Nitrosomonas europaea ATCC 19718. Appl. Environ.

689

Microbiol. 2014, 80, 4930-4935.

690

(61) Yu, R.; Kampschreur, M. J.; van Loosdrecht, M. C. M.; Chandran, K. Mechanisms and

691

specific directionality of autotrophic nitrous oxide and nitric oxide generation during

692

transient anoxia. Environ. Sci. Technol. 2010, 44, 1313-1319.

693

(62) Kozlowski, J. A.; Kits, K. D.; Stein, L. Y. Comparison of nitrogen oxide metabolism

694

among diverse ammonia-oxidizing bacteria. Front. Microbiol. 2016, 7, 1090.

695

(63) Caranto, J. D.; Lancaster, K. M. Nitric oxide is an obligate bacterial nitrification

696

intermediate produced by hydroxylamine oxidoreductase. Proc. Natl. Acad. Sci. U. S. A.

697

2017, 114, 8217-8222.

698

(64) Anderson, J. H. The metabolism of hydroxylamine to nitrite by Nitrosomonas.

699

Biochem. J. 1964, 91, 8-17.

700

(65) Silva, C. M.; Dias, I. C.; Pliego, J. R., Jr. The role of ammonia oxide in the reaction of

701

hydroxylamine with carboxylic esters. Org. Biomol. Chem. 2015, 13, 6217-6224.

702



A

Page 34 of 39

B

5x107

Asulam TP230 TP274

1.0

Peak Area

Asulam C/C0

4x107

0.8 0.6 0.4 Active biomass Medium ctrl Heat-inactivated ctrl

0.2

3x107 2x107 1x107 0

0 0

24

48

72 96 Time (h)

120

144

0

168

24

48

72

96

120

144

168

Time (h)

Figure 1. Biotransformation (A) and TP formation (B) of asulam (initial concentration: 10 µg/L) in the AOB/NOB co-culture (n=3; TP274 and TP230 were not detected in the medium or the heat-inactivated control).


Page 35 of 39


1.2

A

Asulam C/C0

1.0 0.8 0.6 0.4 10 µg/L asulam + 2 mM NH4-N

0.2

10 µg/L asulam + 0.05 mM NH4-N

0

B

Nitrate (mM)

20 15 10

0 µg/L asulam + 2 mM NH4-N

5

10 µg/L asulam + 2 mM NH4-N 10 µg/L asulam + 0.05 mM NH4-N

16S rRNA gene (copies/mL)

0

C

8x108

N. europaea 0 µg/L asulam + 2 mM NH4-N

6x108

10 µg/L asulam + 2 mM NH4-N 10 µg/L asulam + 0.05 mM NH4-N

4x108

2x108

108

0

24

48

72 96 Time (h)

120

144

168

Figure 2. Asulam biotransformation by the AOB/NOB co-culture grown with reamended and limited NH4-N (A: asulam removal; B: nitrate formation; C: cell growth of the AOB N. europaea; 2 mM ammonium was added repeatedly as indicated by arrows; 0.05 mM ammonium was only added at 0 h; n=3).



Page 36 of 39

20 A 0 µg/L asulam 10 µg/L asulam 10 µg/L asulam + octyne

0.8 Asulam C/C0

Nitrate (mM)

15

B

1.0

10

0.6 octyne added 0.4

5 octyne added

10 µg/L asulam 10 µg/L asulam + octyne

0.2

0

0

0 30 60 70

72

74

96 120144168

0 30 60 70

Time (h)

5x108

N. europaea

C

108 octyne added 0 µg/L asulam 10 µg/L asulam 10 µg/L asulam + octyne

2x107 0

24

48

72

96

74

96 120144168

Time (h)



5x108

72

120

Nitrobacter sp.

D

octyne added 108

0 µg/L asulam 10 µg/L asulam 10 µg/L asulam + octyne 107

144

0

168

24

48

72

96

120

144

168

Time (h)

Time (h)

Figure 3. Nitrate formation from a total of 14 mM ammonium addition (nitrite was below detection limit under all conditions) (A), asulam removal (B), growth of N. europaea (C) and Nitrobacter sp. (D) in the AOB/NOB co-culture with and without the addition of octyne (72h results were measured before octyne addition, n=3).


Page 37 of 39


100

** *

Removal (%)

80

*

* *

*

*

* *

60

*

*

NH2OH + fresh medium_12h

*

*

*

40

*

20

AOB/NOB co-culture_168h NH2OH + fresh medium_168h

*

*

* * * *

NO + fresh medium_168h Fresh medium_168h

* *

*

* 0 -20

ne As Fu ula m ro se m Fe nh i de ex a R uf m i d in a Be mid e za fi b C ar r b e a te nd Tr i m az i m et ho pr M im on u C lo ron m a Ac zo ne e L e ta m ip ve ri ti r ac d et am Te Irg bu aro fe no l Th zid e ia cl op rid

n ci

i ti di

ha

an R

et m

do In

M

ia

ns

er in

-40

Figure 4. Comparison of MP transformation in NH2OH- or NO-added sterile fresh medium and in the AOB/NOB co-culture (* indicates significant difference from the 168h removal in the fresh medium control, p < 0.05, n = 3).



1.2x107 1.0x107

AOB/NOB co-culture_168h NH2OH + fresh medium_168h

Peak area

8.0x106 6.0x106 4.0x106 2.0x106 0

TP274

TP230

Figure 5. Formation of asulam TPs in the AOB/NOB co-culture and in the NH2OH-treated fresh medium after a 168-h incubation with the same addition of asulam (neither of the two TPs was detected in the medium or heat-inactivated biomass control, n=3).


Page 38 of 39

Page 39 of 39


TOC art


Ammonia Monooxygenase-Mediated Cometabolic Biotransformation

Recommend Documents