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Impact of the Ammonium Availability on Atenolol Biotransformation during Nitrification Yifeng Xu, Zhiguo Yuan, and Bing-Jie Ni ACS Sustainable Chem. Eng., Just Accepted Manuscript • DOI: 10.1021/acssuschemeng.7b01319 • Publication Date (Web): 02 Jul 2017 Downloaded from http://pubs.acs.org on July 9, 2017

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ACS Sustainable Chemistry & Engineering

1

Impact of the Ammonium Availability on Atenolol Biotransformation during

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Nitrification

3 Yifeng Xu1, Zhiguo Yuan1, Bing-Jie Ni1,2,*

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1

Advanced Water Management Centre, The University of Queensland, St. Lucia, Brisbane, QLD

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4072, Australia 2

Department of Civil Engineering, Monash University, Clayton, VIC3800, Australia

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*Corresponding author: Phone: + 61 7 3346 3219; Fax: +61 7 3365 4726; E-mail: [email protected]

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1

ACS Paragon Plus Environment


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ABSTRACT

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The impact of the ammonium availability on atenolol biotransformation at an environmentally

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relevant level of 15 µg L-1 by enriched nitrifying cultures was investigated in terms of atenolol

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degradation

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concentrations of growth substrate ammonium (0, 25, and 50 mg-N L-1) were applied constantly

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during batch experiments. The results suggested the higher ammonium concentrations led to

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lower atenolol removal rates probably due to the substrate competition between ammonium and

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atenolol. The formation of the biotransformation product atenolol acid was positively related to

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the ammonium oxidation activity, resulting in a higher amount at the end of experiments at

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higher ammonium concentrations. Linear correlations between ammonia oxidation rate and

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atenolol degradation rate at ammonium levels of 25 and 50 mg-N L-1 suggested the

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cometabolism of atenolol by ammonia oxidizing bacteria (AOB) in the presence of ammonium.

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The biotransformation reaction, i.e., hydroxylation on amide group to carboxylic group, could be

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catalyzed by the non-specific ammonia monooxygenase (AMO) of AOB. Comparison between

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atenolol degradation at ammonium levels of 0 and 50 mg-N L-1 demonstrated the formation of

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atenolol acid was independent of the ammonium availability. This work might give further

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implication in how to prevent pharmaceuticals from entering into the environment.

kinetics

and

biotransformation

product

formation

dynamics.

Different

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Keywords: Ammonia oxidizing bacteria; ammonium; atenolol; biotransformation product;

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cometabolism; biodegradation

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2


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For Table of Contents Use Only

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Synopsis: Atenolol removal efficiency was adversely associated with ammonium availability

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while atenolol acid formation was positively linked to ammonia oxidizing activity.

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INTRODUCTION

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For recent decades emerging organic micropollutants including pesticides, pharmaceuticals and

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personal care products have attracted increasing public and research concerns due to their

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potential for adverse effects on human and ecosystems.1-3 Their ubiquitous occurrence has been

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reported as ranging from a few nano-gram per litre to several hundred micro-gram per litre in

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wastewater, surface water and groundwater.4 Wastewater treatment processes have been

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identified as the main point of discharge of pharmaceuticals into the environment as they were

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originally designed for oxygen demand removal.5,6 In addition, biodegradation of

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pharmaceuticals in wastewater treatment processes might form transformation products that are

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potentially more toxic and persistent in the aquatic environment.7 Thus, an in-depth

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understanding of the fate of pharmaceuticals in wastewater treatment is needed to assess the

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removal of these compounds and their biotransformation products.

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The positive relationships between pharmaceutical biotransformation and nitrification rates have

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been reported in previous studies.8,9 Enhanced pharmaceutical removal were found to be

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attributed to the non-specific enzyme ammonia monooxygenase (AMO) from ammonia

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oxidizing bacteria (AOB), with a broad substrate range including aromatic and aliphatic

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compounds.10-12 It was hypothesized that some pharmaceuticals could be biotransformed

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cometabolically by AOB in nitrifying cultures.13,14 In particular, higher biodegradation

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efficiencies of the emerging micropollutants were obtained at higher nitrogen loading rates that

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could promote the development of biomass with high nitrification activities.11 As the obligatory

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growth substrate for AOB, investigation on the relationship between ammonium and

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pharmaceuticals would give an insight in enhancing their removal, therefore preventing their

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input into the environment. The influence of the initial ammonium concentration on

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biotransformation of ibuprofen has been investigated in previous reports.15 However, the effect

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of the constant presence of ammonium on pharmaceutical biodegradation and on transformation

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product formation has not been studied yet for AOB induced cometabolism in enriched nitrifying

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cultures.

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Atenolol was among the most frequently reported pharmaceuticals in the wastewater with the

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highest concentration being up to 25 µg L-1.16 Biodegradation of atenolol was linked to the

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activity of AOB and heterotrophs in the enriched nitrifying cultures.17 Our recent work mainly

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focused on the identification and elucidation of biotransformation products and pathways of

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atenolol at high initial concentration (15 mg L-1) under different metabolic conditions.18 We

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determined that the additional research should be conducted at an environmentally relevant

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concentration. Therefore, the objective of this study was to investigate the impact of the

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ammonium availability on atenolol biotransformation at 15 µg L-1 by an enriched nitrifying

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sludge. Batch experiments were conducted by holding the ammonium concentrations (0, 25, and

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50 mg-N L-1) constant for 240 h to evaluate the atenolol degradation kinetics and the

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biotransformation product formation dynamics.

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MATERIALS AND METHODS

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Chemicals. Atenolol (≥98%), atenolol acid, allylthiourea (ATU, 98%), the internal standard

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atenolol-d7 (≥97%) and all the other organic solvents (LC grade) were purchased from Sigma-

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Aldrich, Australia. The standard stock solution of atenolol was prepared on a weight basis in

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methanol at 1 g L-1 and kept frozen at -20 °C. Working standards used for calibration curves

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were prepared by appropriate dilution of the standard stock solution with purified Milli-Q water

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(Millipore, Inc.). Atenolol feed solution at 1 mg L-1 was obtained by dissolving the atenolol

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powder in Milli-Q water, which was used to provide the initial atenolol concentration for all the

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batch biodegradation experiments.

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Nitrifying culture enrichment. Seed sludge for the nitrifying enrichment was collected from a

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domestic wastewater treatment plant in Brisbane, Australia. The enrichment process was

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achieved in the lab-scale sequencing batch reactor (SBR) at a volume of 8 L. It was operated on

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a 6-h cycle (260 min fill, 30 min aerobic react, 1 min waste, 60 min settle, 9 min decant) during

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the time period when enriched nitrifying biomass was collected for batch experiments with

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atenolol. For each cycle, 2 L feed solution was dosed into SBR, which consisted of 5.63 g of

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NH4HCO3 (1 g NH4+-N), 5.99 g of NaHCO3, 0.064 g of each of KH2PO4 and K2HPO4 and 2 mL

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of a trace element solution per litre.19 The trace element stock solution contained: 1.25 g L-1

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EDTA, 0.55 g L-1 ZnSO4·7H2O, 0.40 g L-1 CoCl2·6H2O, 1.275 g L-1 MnCl2·4H2O, 0.40 g L-1

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CuSO4·5H2O, 0.05 g L-1 Na2MoO4·2H2O, 1.375 g L-1 CaCl2·2H2O, 1.25 g L-1 FeCl3·6H2O and

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44.4 g L-1 MgSO4·7H2O. Reactor pH and dissolved oxygen (DO) were continuously monitored

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using miniCHEM meters and controlled in the range of 7.5-8.0 and 2.5-3.0 mg L-1, respectively,

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with programmed logic controllers (PLC). Hydraulic retention time (HRT) and solid retention

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time (SRT) were 24 h and 15 d, respectively.

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Due to the relatively slow growth rate of nitrifying bacteria, the SBR was operated for more than

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one year until the nitrifying culture was successfully enriched in the SBR with stable

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performance at steady state. The enriched nitrifying culture was then taken from the SBR and

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utilized for batch biodegradation experiments. The bacterial community analysis was conducted

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to obtain the fractions of enrolled bacteria using fluorescence in situ hybridization (FISH).20

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AOB and nitrite oxidizing bacteria (NOB) together would contribute more than 80% of the

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enriched nitrifying biomass with their corresponding percentages of 46 ± 6% (n=20) and 38 ±

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5% (n=20), respectively.

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Batch experiments with different ammonium availability. Covered by the aluminum foil, the

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beakers were used to carry out the biodegradation experiments. 4-L beakers were filled with

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enriched nitrifying cultures taken from the SBR. The Mixed liquor volatile suspended solids

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(MLVSS) in the beakers was achieved at approximately 1000 mg L-1 by diluting the enriched

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nitrifying cultures with the SBR effluent at the beginning of the experiments. The target

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pharmaceutical, atenolol, was then added into the beakers to obtain an initial environmentally

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relevant concentration (15 µg L-1). Different ammonium concentrations (0, 25 and 50 mg L-1)

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were controlled constantly for each biodegradation experiment with different ammonia oxidation

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levels in duplicates (n=2). Ammonium bicarbonate was supplied as the growth substrate at the

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beginning of the experiments with ammonium addition and was further frequently dosed into the

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beakers with sodium bicarbonate as a pH adjustment method as well as maintaining a constant

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level of ammonium. The experiments without ammonia oxidation were conducted in duplicates

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where ammonium concentration was zero during the whole time period. The control experiments

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were carried out to study the abiotic and hydrolytic degradation of atenolol with autoclaved

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biomass (121 °C and 103 kPa for 30 min)21 and without any biomass, respectively. For all batch

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experiments, pH was controlled between 7.5 and 8.0 and DO was supplied by aeration in the

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range of 2.5 and 3.0. After mixing well, the samples were taken periodically each day for 240 h

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for chemical analysis of atenolol and its products.

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Chemical analysis. Mixed liquor suspended solids (MLSS) and the volatile fraction (MLVSS)

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were measured at 0, 120 and 240 h during the batch biodegradation experiments according to the

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standard methods.22 NH4+-N concentrations were measured every hour for the first 24 h and

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every two hours for the following experimental period using a Lachat QuikChem8000 Flow

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Injection Analyzer (Lachat Instrument, Milwaukee). Figure S1 in Supporting Information (SI)

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demonstrated that the ammonium concentrations were controlled nearly constant during the

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entire experimental period for batch biodegradation experiments in the presence of ammonium.

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As illustrated in Figure S1, no ammonium was detected in the experiments in the absence of

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ammonium. In addition, nitrite was not accumulated for all the batch experiments (less than 1 mg

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L-1), which indicated the absence of nitration reactions.23 Furthermore, the nitrate concentration

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was similar to that in the SBR effluent (up to 1000 mg L-1).

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For sample preparation, solid phase extraction (SPE) with vacuum manifold (J. T. Baker, The

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Netherlands) was carried out first in order to achieve the pre-concentration of atenolol and its

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products at lower levels. Oasis HLB cartridges (6 mL, 200 mg, Waters, USA) were conditioned

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using 10 mL methanol and 10 mL Milli-Q water.24 50 mL samples were centrifuged at 3200 g

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for 5 min and then the supernatants were passed through the pre-conditioned cartridges at a flow

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rate of approximately 5 mL min-1. Following sample pre-concentration, cartridges were rinsed

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with 5 mL Milli-Q water and were dried under vacuum for 30 min. The analytes were eluted

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with 10 mL methanol and 10 mL of hexane/acetone (50:50, v/v) at a flow rate of 1 mL min-1.

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Sample extracts were evaporated to dryness using a gentle stream of nitrogen and were

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reconstituted with 250 µL methanol and 750 µL Mmilli-Q water. 20 µL atenolol-d7 was added

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into each sample as an internal standard to achieve a concentration of 50 µg L-1 before further

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analysis.

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The liquid chromatograph used in this study was an ultra-fast liquid chromatography (UFLC)

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(Shimadzu, Japan). An Alltima C18 column (Alltech Associates Inc., USA) was used for

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chromatographic separation using Milli-Q (A) and acetonitrile (B) as mobile phases at a flow

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rate of 1 mL min-1. The gradient elution program was set as follows: initial, 0% B; 0-0.5 min, 0-

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5% B; 0.5-13 min, 5-20% B; 13-18 min, 20-50% B; 18-20 min, 50-100% B; 20-24 min, 100% B;

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24-25 min, 100-5% B. The total running time including the conditioning of the column to the

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initial conditions was 27 min. The column oven was set to 40 °C. The injection volume was 20

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µL. The mass spectrometric analysis was performed on a 4000 QTRAP hybrid triple quadruple-

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linear ion trap mass spectrometer (QqLIT-MS) equipped with a Turbo Ion Spray source (Applied

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Biosystems-Sciex, USA) under positive ionization mode (ESI+) for all acquisitions. The drying

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gas (50 psi) was used at 500 °C with curtain gas 30 psi and spraying gas 50 psi applied. For

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qualitative purposes, full scan mode was performed on the samples at the declustering potential

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of 80 V and mass range of 50-500 amu followed by product ion scan mode and sequential

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fragmentation. For quantitative purposes, multiple reaction monitoring (MRM) mode was

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conducted with two MRM transition ions for each compound (Table 1), the first one for

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quantification and the second one for confirmation of the compound. The limit of detection and

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limit of quantification for atenolol and atenolol acid were 0.020 and 0.067 µg L-1, 0.013 and

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0.043 µg L-1,respectively.

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RESULTS

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Nitrification performance and control experiments. Prior to the beginning of the batch

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biodegradation experiments with atenolol, the MLVSS concentration of SBR was stable at

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1437.6±112.9 mg L-1 (mean and standard errors, respectively, n=10). The stable nitrification

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performance was achieved in the SBR, which was assessed through a series of cycle studies in

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duplicates (n=2): an initial pulse of 20 mg L-1 ammonium was added into the enriched nitrifying

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sludge at the MLVSS concentration of 95.0±6.7 mg L-1. The concentration profiles of different

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nitrogen species were shown in Figure 1A as NH4+-N, NO2--N and NO3--N, respectively.

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Ammonium was completely consumed by the enriched nitrifying sludge within 5 h, leading to a

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sharp increase of NO3--N to approximately same concentration of the initial NH4+-N level. The

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concentration of nitrite was increased at first 4 h and then decreased to zero until the end of

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experiments, which indicated its consumption by NOB. Overall, a good nitrification performance

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of the enriched nitrifying sludge was achieved with the highest ammonia oxidizing rate

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(approximately 52 mg NH4+-N gVSS-1 h-1).

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The abiotic control and hydrolytic control experiments demonstrated a rather stable atenolol

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concentration over 240 h without the formation of any transformation products (Figure S2).

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Given the lower adsorption ability and the aluminum cover avoiding photodegradation,

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biodegradation by the enriched nitrifying biomass was therefore determined to be the main

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removal mechanism in the investigated batch experiments.

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Impact of ammonium availability on atenolol biodegradation. To determine the effect of the

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growth substrate (ammonium) availability on the cometabolic activity on atenolol by the

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enriched nitrifying culture, a series of ammonium concentrations among 0, 25 and 50 mg L-1

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were applied in different batch biodegradation experiments with initial 15 µg L-1 atenolol. It can

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be concluded from Figure 2A that the removal efficiencies of atenolol decreased with the

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increase of the availability ammonium at constant levels during the experimental period. Such

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observation is different from the previous studies where the degradation of selected

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pharmaceuticals (e.g., clofibric acid, diclofenac, carbamazepine, propyphenazone) was enhanced

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at higher initial ammonium concentration.11 In comparison, the ammonium concentration was

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provided with pulse feeding at the beginning of the experiments in previous studies, leading to

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completely different ammonium availability (constant concentrations in this study) during the

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time course. Enriched nitrifying sludge responsible for the cometabolic biodegradation was

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reported to have a different affinity for each pharmaceutical compounds, probably due to a

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preferential substrate selection to AMO active sites.12 Ammonia oxidation rate was calculated

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based on the measured ammonium concentration and the adding volume of ammonium

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bicarbonate for each time. Based on Figure 1B, the ammonium availability at constant level had

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a positive effect on ammonia oxidation rate, whereas atenolol removal efficiency was adversely

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related to the constant ammonium concentration.

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Impact of ammonium availability on formation of atenolol acid. In this study, the influence

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of ammonium availability on the formation of biotransformation products of atenolol was also

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investigated (Figure 2B). Accompanied with the decrease of atenolol (Figure 2A), only one

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biotransformation product was found in all the experiments and was confirmed as atenolol acid.

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The structural elucidation of atenolol acid was detailed in previous studies.25 This was different

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from our previous report where four products were identified in the presence of ammonium,

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probably due to the applied higher initial concentration of atenolol.18 However, as a major

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biotransformation product of atenolol, understanding the formation dynamics of atenolol acid

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was necessary to reflect the impact of ammonium availability. It can be clearly concluded that

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atenolol acid was produced more in the presence of ammonium than that in the absence of

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ammonium, and was formed increasingly with the increase of ammonium availability. Although

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AOB were able to degrade pharmaceuticals during ammonium starvation,26,27 its contribution to

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the formation of atenolol acid was yet less than AOB with the adequate growth substrates. It was

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also confirmed that the formation of biotransformation products in the presence of ammonium

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was due to cometabolic biodegradation by AOB, as ammonia oxidation rate showed a positive

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correlation with the formation efficiency of atenolol acid. The biotransformation of atenolol to

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atenolol acid was via hydrolysis on the primary amide group to the carboxylic acids catalyzed by

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AMO, which was dependent on the nitrifying activities.25,28

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Comparison between atenolol biodegradation with and without ammonium. As shown in

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Figure 3A, a gradual decrease of atenolol at an initial concentration of 15 µg L-1 was found

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during the entire experimental period in the presence of ammonium (at 50 mg L-1). The overall

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removal rate of atenolol could reach 88.0% with the final concentration of 2.2 µg L-1 in the batch

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reactor. The formation of atenolol acid was correlated to the decrease of atenolol. Atenolol acid

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experienced a sharp increase for first 120 h when ammonia oxidation rate was higher. With the

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available reference standard, atenolol acid was quantified to increase to 15.4 µg L-1. Furthermore,

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atenolol acid was the major biotransformation product in the presence of ammonium through the

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nearly constant mass balance analysis. Figure 3B plotted the concentration profiles of atenolol

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and atenolol acid in the absence of ammonium (at 0 mg L-1). Dropping from the initial 15 µg L-1

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concentration, atenolol experienced a sharp decrease to 51.5% in the first 24 h. The degradation

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of atenolol became slower with the decreasing ammonia oxidation rate until the atenolol

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concentration became 0.6 µg L-1. The concentration of atenolol acid was increased to 6.8 µg L-1

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at 240 h with an approximate conversion rate of 29.1% from atenolol biodegradation. In our

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previous research on atenolol acid biodegradation experiments at an initial concentration of 15

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mg L-1, three biotransformation products were identified in the presence of ammonium at a low

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conversion rate.18 One structure-unknown product with the nominal mass of 227 was also

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proposed to be formed from atenolol acid biodegradation in the absence of ammonium.18

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Decreasing mass balance results in Figure 3B indicated the possible presence of other

245

transformation products, which were not identified probably due to the lower concentration than

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the detection limit. This requires further investigation. The removal efficiency of atenolol was

12


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higher in the absence of ammonium (97.4%) than that in the presence of ammonium, which was

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contradictory to the results reported previously.13 Cometabolism would not contribute to atenolol

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biodegradation in the absence of ammonium significantly due to minor ammonium concentration

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released from cell lysis during bacterial decay. Heterotrophs were also capable to transform

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atenolol to atenolol acid, confirmed in the experiments with addition of ATU previously.17,18 The

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contribution of heterotrophs to pharmaceutical biodegradation in mixed activated sludge

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communities might occur via metabolic or cometabolic pathways.10,13 In this study, the

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cometabolic biodegradation induced by AOB in the enriched nitrifying cultures in the presence

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of ammonium showed the lower contribution to atenolol removal at the environmentally relevant

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concentration compared to the biodegradation by nitrifying culture in the absence of ammonium,

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which could be caused by the substrate competition especially in the case when growth substrate

258

concentration was higher.15

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Kinetic analysis of atenolol biodegradation. An exponential decrease of atenolol concentration

260

was observed for each experiment without the presence of ammonium, and in the presence of 25

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mg L-1 and 50 mg L-1 ammonium (NH4+-N) (Figure 2A), indicating a pseudo first order

262

degradation kinetics for atenolol removal by the enriched nitrifying sludge (Figure S3). The

263

biodegradation rate constants kbio could be calculated using:

264

(1)

265

where C is the total compound concentration (µg L-1), C0 is the measured initial concentration

266

(µg L-1), t is time (h), kbiol is the biodegradation rate constant (L gVSS-1 h-1), XVSS is the VSS

267

concentration in the batch beakers (gVSS L-1) and t1/2 is the biodegradation half-life (h).

268

For the batch experiments without the presence of ammonium, the biodegradation rate constant

269

of atenolol was calculated as 0.018±0.001 L gVSS-1 h-1 with the degradation half-life of 40.2 h.

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For atenolol biodegradation in the presence of ammonium, the half-lives were 63.1 h and 77.3 h

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at the constant ammonium concentrations of 25 and 50 mg L-1, respectively. The degradation

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constants of 0.0078±0.0004 and 0.0072±0.0004 L gVSS-1 h-1 indicated a slower degradation of

273

atenolol in the presence of ammonium compared to the batch experiments without the presence

274

of ammonium. This further supported the notion that atenolol could be degraded during

275

ammonium starvation like other pharmaceuticals.15 The substrate competition existed between

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ammonium and atenolol might lead to a decreasing degradation rate under high concentrations of

277

ammonium.10 These degradation constants were lower than the reported values for atenolol (1.1-

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1.9 L gSS-1 d-1),29 likely due to the unaccustomed sludge to atenolol in this work.

279 280

DISCUSSION

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Competitive inhibition of ammonium on atenolol biodegradation. As shown in Figure 2A,

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the removal efficiency of atenolol was adversely linked to the ammonium concentration. The

283

higher removal of atenolol was observed under the lower ammonium concentration at constant

284

level. Such observations suggested there might exist a competition between the growth substrate

285

(ammonium) and cometabolic substrate (atenolol), in particular when the concentration of

286

growth substrate was much higher than that of the non-growth substrate,12 which was previously

287

confirmed in the biodegradation study on cis-1,2-dichloroethene and tetrachloroethylene.30,31 As

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the ammonia oxidation and hydroxylation of atenolol were both catalyzed by AMO, it could be

289

hypothesized that the competition for AMO active sites between two substrates would be the

290

crucial mechanism in this study and ammonia oxidation might be the limiting step, leading to the

291

lower removal rate of atenolol under higher constant ammonium concentrations. When

292

ammonium concentrations were maintained at zero in the batch experiments, metabolism

14


Page 15 of 28

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293

induced by enriched nitrifying sludge might be the main degradation mechanism. There was

294

evidence that AOB can transform pharmaceuticals during ammonium starvation.15,26 For

295

example, ibuprofen would not be degraded unless ammonia was depleted completely, therefore

296

showing a decreased biodegradation rate at high ammonium concentration (100 mg L-1).15

297

However, the formation of atenolol acid was enhanced with increasing ammonium concentration

298

(Figure 2B), specifically with increasing ammonium oxidizing rates shown in Figure 1B. AMO

299

activities might be the dominant factor determining the formation of atenolol acid as AMO could

300

catalyze hydroxylation and oxidation.32,33 Therefore, the final concentrations of atenolol acid

301

were increased from 8.6 to 15.4 µg L-1 with the increasing ammonium concentrations from 25 to

302

50 mg L-1. Due to the competitive inhibition by atenolol,17 ammonia oxidation rates also showed

303

a decreasing trend along with the time course thus leading to a slow increase of atenolol acid in

304

both cases. The formation efficiency of atenolol acid was the lowest without the presence of

305

ammonium amongst these experiments, probably due to the lowest AMO activities lacking of

306

growth substrates.

307

Relationship between ammonia oxidation and atenolol degradation. The cometabolic

308

biodegradation by AOB was also confirmed from the linear positive relationship between

309

ammonia oxidation rate and atenolol biodegradation rate as shown in Figure 4A, which was also

310

reported in previous literature.34 This linear relationship was valid at the investigated mole ratio

311

of atenolol to ammonium from 2.4×10-6 to 2.1×10-5 in the biodegradation experiment with 50 mg

312

L-1 ammonium while the corresponding ratio was from 2.8×10-6 to 3.6×10-5 in the biodegradation

313

experiments with 25 mg L-1 ammonium. The atenolol biodegradation rate decreased from

314

7.6×10-4 to 3.3×10-5 µmol gVSS-1 h-1 whereas the ammonia oxidation rate decreased from 2.8 to

315

0.02 mmol NH4+-N gVSS-1 h-1 along with the experimental time at the constant 25 mg L-1

15



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Page 16 of 28

316

ammonium. For batch experiments at the constant 50 mg L-1 ammonium, the atenolol

317

degradation rate decreased from 7.8×10-4 to 5.6×10-5 µmol gVSS-1 h-1 with a decreasing ammonia

318

oxidation rate from 4.2 to 0.2 mmol NH4+-N gVSS-1 h-1. Comparing the fitted slopes between the

319

experiments at constant 25 mg L-1 and 50 mg L-1 ammonium, it could be clearly concluded that a

320

higher atenolol degradation rate would be achieved at the lower concentration of growth

321

substrate with the same ammonia oxidation rate, further supporting the substrate competition for

322

AMO active sites between ammonium and atenolol.13,35

323

The observed relationship between ammonia oxidation rate and atenolol acid formation rate in

324

Figure 4B indicated that the formation of biotransformation product of atenolol was consistent

325

with the monooxygenase activity.36 This suggested the involvement of AMO in the

326

biotransformation reaction.33 When ammonia oxidation rate was lower than 14.5 mg NH4+-N

327

gVSS-1 h-1, the higher formation rate of atenolol acid would be achieved at the higher ammonium

328

concentrations provided that the same ammonia oxidation rate was obtained for both constant

329

ammonium concentrations conditions (25 or 50 mg L-1) (Figure 4B). The relationship implied

330

that the batch experiments at 25 mg L-1 ammonium would show a higher formation rate of

331

atenolol acid at the assumed same ammonia oxidation rate when it was higher than 14.5 mg

332

NH4+-N gVSS-1 h-1. This might be due to the involvement of cometabolism and substrate

333

competition together.10,13 The competition might play an important role in formation of

334

biotransformation products when ammonia oxidizing rate was higher than a critical value (e.g.

335

14.5 mg NH4+-N gVSS-1 h-1 in this study), which requires further confirmation.

336

With regards to the corresponding concentration profiles of atenolol acid in Figure 2B, it

337

demonstrated a relatively rapid increasing trend under the lower ammonium concentration (25

338

mg L-1) for the first 48 h. However, ammonia oxidation rates under the higher ammonium

16


Page 17 of 28

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339

concentration (50 mg L-1) were still higher than those under the lower ammonium concentration

340

(25 mg L-1) for each sampling time until 240 h, thus leading to a higher atenolol acid formation

341

rate correspondingly and a higher final concentration of atenolol acid.

342

Atenolol biodegradation pathway with different ammonium availability. Different metabolic

343

conditions with regards to the presence or the absence of ammonium were applied in this study

344

to investigate the degradation of atenolol and the formation of its biotransformation product

345

atenolol acid by the enriched nitrifying cultures. It was shown that atenolol could be

346

hydroxylated to its carboxylic form, i.e., atenolol acid, regardless of the presence/availability of

347

the growth substrate ammonium. Atenolol acid was also a biotransformation product observed in

348

the conventional activated sludge and membrane bioreactor as well as activated sludge receiving

349

sanitary sewage.25,37 This biotransformation reaction was previously reported to be catalyzed by

350

amidohydrolase produced from bacteria.38 AMO was a non-specific enzyme with a broad

351

substrate range.32,39 This cometabolism-induced biotransformation reaction was catalyzed by

352

AMO, which was mostly produced by AOB in the presence of ammonium. In our previous work

353

on atenolol biodegradation at initial concentration of 15 mg L-1, microbial induced hydroxylation

354

on the amide group to its carboxylic moiety produced the biotransformation product, atenolol

355

acid, which is consistent with this study.18 However, 1-isopropylamino-2-propanol, 1-amino-3-

356

phenoxy-2-propanol and an unknown product with the nominal mass of 227 were also identified

357

under higher initial concentration of atenolol compared to this study. The previous reports on the

358

types of biotransformation products formed at different initial concentrations were contradictory.

359

For examples, the same metabolites and degradation routes were reported for trimethoprim at

360

initial 20 mg L-1 and 20 µg L-1 while different biotransformation products of trimethoprim were

361

identified in nitrifying activated sludge under different initial concentrations (500 µg L-1and 5 µg

17



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Page 18 of 28

362

L-1).40,41 This might be caused by the types of nitrifying activated sludge applied in different

363

research. Given the identical nitrifying biomass utilized in our previous work and in this study,18

364

the negative identification of another three biotransformation products might be due to the lower

365

initial concentration of atenolol (15 µg L-1) in this study. The reported lower conversion rates

366

from atenolol to 1-isopropylamino-2-propanol and 1-amino-3-phenoxy-2-propanol might result

367

in the lower concentration of possible products than the limit of detection, which needs more

368

efforts in the future research. Nevertheless, the findings in this work might give an implication on

369

how to prevent pharmaceuticals entering into the environment by concomitant enhanced removal

370

of ammonium and pharmaceuticals through optimization of existing wastewater treatment

371

processes.

372 373

ASSOCIATED CONTENT

374

Supporting Information

375

This file contains ammonium concentration profiles during the biodegradation experiments,

376

atenolol concentration profiles in the control experiments and the linear regressions for atenolol

377

biodegradation at different ammonium concentrations.

378 379

ACKNOWLEDGEMENT

380

This study was supported by the Australian Research Council (ARC) through Discovery Early

381

Career Researcher Award DE130100451. Dr. Bing-Jie Ni acknowledges the support of

382

Australian Research Council Future Fellowship FT160100195.

383 384

REFERENCES

18


Page 19 of 28

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385 386 387

1. Luo, Y.; Guo, W.; Ngo, H. H.; Nghiem, L. D.; Hai, F. I.; Zhang, J.; Liang, S.; Wang, X. C., A review on the occurrence of micropollutants in the aquatic environment and their fate and removal during wastewater treatment. Sci. Total Environ. 2014, 473-474, 619-641.

388 389

2. Fenner, K.; Canonica, S.; Wackett, L. P.; Elsner, M., Evaluating pesticide degradation in the environment: Blind spots and emerging opportunities. Science 2013, 341 (6147), 752-758.

390 391 392

3. Ternes, T.; Joss, A.; Oehlmann, J., Occurrence, fate, removal and assessment of emerging contaminants in water in the water cycle (from wastewater to drinking water). Water Res. 2015, 72, 1-2.

393 394 395

4. Petrie, B.; Barden, R.; Kasprzyk-Hordern, B., A review on emerging contaminants in wastewaters and the environment: Current knowledge, understudied areas and recommendations for future monitoring. Water Res. 2015, 72, 3-27.

396 397 398 399

5. Joss, A.; Zabczynski, S.; Göbel, A.; Hoffmann, B.; Löffler, D.; McArdell, C. S.; Ternes, T. A.; Thomsen, A.; Siegrist, H., Biological degradation of pharmaceuticals in municipal wastewater treatment: Proposing a classification scheme. Water Res. 2006, 40 (8), 16861696.

400 401 402 403

6. Sengupta, A.; Lyons, J. M.; Smith, D. J.; Drewes, J. E.; Snyder, S. A.; Heil, A.; Maruya, K. A., The occurrence and fate of chemicals of emerging concern in coastal urban rivers receiving discharge of treated municipal wastewater effluent. Environ. Toxicol. Chem. 2014, 33 (2), 350-358.

404 405 406

7. Ternes, T. A.; Bonerz, M.; Herrmann, N.; Teiser, B.; Andersen, H. R., Irrigation of treated wastewater in Braunschweig, Germany: An option to remove pharmaceuticals and musk fragrances. Chemosphere 2007, 66 (5), 894-904.

407 408

8. Batt, A. L.; Kim, S.; Aga, D. S., Enhanced biodegradation of iopromide and trimethoprim in nitrifying activated sludge. Environ. Sci. Technol. 2006, 40 (23), 7367-7373.

409 410 411

9. Margot, J.; Lochmatter, S.; Barry, D. A.; Holliger, C., Role of ammonia-oxidizing bacteria in micropollutant removal from wastewater with aerobic granular sludge. Water Sci. Technol. 2016, 73 (3), 564-575.

412 413 414

10. Fischer, K.; Majewsky, M., Cometabolic degradation of organic wastewater micropollutants by activated sludge and sludge-inherent microorganisms. Appl. Microbiol. Biotechnol. 2014, 98 (15), 6583-6597.

415 416 417

11. Tran, N. H.; Urase, T.; Kusakabe, O., The characteristics of enriched nitrifier culture in the degradation of selected pharmaceutically active compounds. J. Hazard. Mater. 2009, 171 (1– 3), 1051-1057.

418 419 420

12. Fernandez-Fontaina, E.; Omil, F.; Lema, J. M.; Carballa, M., Influence of nitrifying conditions on the biodegradation and sorption of emerging micropollutants. Water Res. 2012, 46 (16), 5434-5444.

421 422 423

13. Tran, N. H.; Urase, T.; Ngo, H. H.; Hu, J.; Ong, S. L., Insight into metabolic and cometabolic activities of autotrophic and heterotrophic microorganisms in the biodegradation of emerging trace organic contaminants. Bioresour. Technol. 2013, 146, 721-731.

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14. Roh, H.; Subramanya, N.; Zhao, F.; Yu, C.-P.; Sandt, J.; Chu, K.-H., Biodegradation potential of wastewater micropollutants by ammonia-oxidizing bacteria. Chemosphere 2009, 77 (8), 1084-1089.

427 428 429

15. Dawas-Massalha, A.; Gur-Reznik, S.; Lerman, S.; Sabbah, I.; Dosoretz, C. G., Co-metabolic oxidation of pharmaceutical compounds by a nitrifying bacterial enrichment. Bioresour. Technol. 2014, 167, 336-342.

430 431 432

16. Verlicchi, P.; Al Aukidy, M.; Zambello, E., Occurrence of pharmaceutical compounds in urban wastewater: Removal, mass load and environmental risk after a secondary treatment— A review. Sci. Total Environ. 2012, 429, 123-155.

433 434 435

17. Sathyamoorthy, S.; Chandran, K.; Ramsburg, C. A., Biodegradation and cometabolic modeling of selected beta blockers during ammonia oxidation. Environ. Sci. Technol. 2013, 47 (22), 12835-12843.

436 437

18. Xu, Y.; Radjenovic, J.; Yuan, Z.; Ni, B. J., Biodegradation of atenolol by an enriched nitrifying sludge: Products and pathways. Chem. Eng. J. 2017, 312, 351-359.

438 439

19. Kuai, L.; Verstraete, W., Ammonium removal by the oxygen-limited autotrophic nitrification- denitrification system. Appl. Environ. Microbiol. 1998, 64 (11), 4500-4506.

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20. Law, Y.; Lant, P.; Yuan, Z., The effect of pH on N2O production under aerobic conditions in a partial nitritation system. Water Res. 2011, 45 (18), 5934-5944.

442 443 444

21. Kassotaki, E.; Buttiglieri, G.; Ferrando-Climent, L.; Rodriguez-Roda, I.; Pijuan, M., Enhanced sulfamethoxazole degradation through ammonia oxidizing bacteria co-metabolism and fate of transformation products. Water Res. 2016, 94, 111-119.

445 446 447

22. APHA. Standard methods for the examination of water and wastewater; American Public Health Association, American Water Works Association, Water Environment Federation: Washington, DC, 1998.

448 449 450

23. Gaulke, L. S.; Strand, S. E.; Kalhorn, T. F.; Stensel, H. D., 17α-ethinylestradiol transformation via abiotic nitration in the presence of ammonia oxidizing bacteria. Environ. Sci. Technol. 2008, 42 (20), 7622-7627.

451 452

24. Xu, Y.; Yuan, Z.; Ni, B.-J., Biotransformation of acyclovir by an enriched nitrifying culture. Chemosphere 2017, 170, 25-32.

453 454 455 456

25. Radjenović, J.; Pérez, S.; Petrović, M.; Barceló, D., Identification and structural characterization of biodegradation products of atenolol and glibenclamide by liquid chromatography coupled to hybrid quadrupole time-of-flight and quadrupole ion trap mass spectrometry. J. Chromatogr. A 2008, 1210 (2), 142-153.

457 458 459

26. Forrez, I.; Carballa, M.; Boon, N.; Verstraete, W., Biological removal of 17α-ethinylestradiol (EE2) in an aerated nitrifying fixed bed reactor during ammonium starvation. J. Chem. Technol. Biotechnol. 2009, 84 (1), 119-125.

460 461 462 463

27. Khunjar, W. O.; Mackintosh, S. A.; Skotnicka-Pitak, J.; Baik, S.; Aga, D. S.; Love, N. G., Elucidating the relative roles of ammonia oxidizing and heterotrophic bacteria during the biotransformation of 17α-ethinylestradiol and trimethoprim. Environ. Sci. Technol. 2011, 45 (8), 3605-3612.

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28. Helbling, D. E.; Johnson, D. R.; Honti, M.; Fenner, K., Micropollutant biotransformation kinetics associate with WWTP process parameters and microbial community characteristics. Environ. Sci. Technol. 2012, 46 (19), 10579-10588.

467 468 469

29. Pomiès, M.; Choubert, J. M.; Wisniewski, C.; Coquery, M., Modelling of micropollutant removal in biological wastewater treatments: A review. Sci. Total Environ. 2013, 443, 733748.

470 471 472

30. Schäfer, A.; Bouwer, E. J., Toluene induced cometabolism of cis-1,2-dichloroethylene and vinyl chloride under conditions expected downgradient of a permeable Fe(0) barrier. Water Res. 2000, 34 (13), 3391-3399.

473 474 475

31. Tsien, H. C.; Brusseau, G. A.; Hanson, R. S.; Wackett, L. P., Biodegradation of trichloroethylene by Methylosinus trichosporium OB3b. Appl. Environ. Microbiol. 1989, 55 (12), 3155-3161.

476 477 478

32. Keener, W. K.; Arp, D. J., Kinetic studies of ammonia monooxygenase inhibition in Nitrosomonas europaea by hydrocarbons and halogenated hydrocarbons in an optimized whole- cell assay. Appl. Environ. Microbiol. 1993, 59 (8), 2501-2510.

479 480 481 482

33. Men, Y.; Han, P.; Helbling, D. E.; Jehmlich, N.; Herbold, C.; Gulde, R.; Onnis-Hayden, A.; Gu, A. Z.; Johnson, D. R.; Wagner, M.; Fenner, K., Biotransformation of two pharmaceuticals by the ammonia-oxidizing archaeon Nitrososphaera gargensis. Environ. Sci. Technol. 2016, 50 (9), 4682-4692.

483 484

34. Yi, T.; Harper Jr, W. F., The link between nitrification and biotransformation of 17αethinylestradiol. Environ. Sci. Technol. 2007, 41 (12), 4311-4316.

485 486

35. Arp, D.; Yeager, C.; Hyman, M., Molecular and cellular fundamentals of aerobic cometabolism of trichloroethylene. Biodegradation 2001, 12 (2), 81-103.

487 488 489

36. Fernandez-Fontaina, E.; Gomes, I. B.; Aga, D. S.; Omil, F.; Lema, J. M.; Carballa, M., Biotransformation of pharmaceuticals under nitrification, nitratation and heterotrophic conditions. Sci. Total Environ. 2016, 541, 1439-1447.

490 491 492

37. Helbling, D. E.; Hollender, J.; Kohler, H. P. E.; Singer, H.; Fenner, K., High-throughput identification of microbial transformation products of organic micropollutants. Environ. Sci. Technol. 2010, 44 (17), 6621-6627.

493 494 495 496

38. Golan-Rozen, N.; Seiwert, B.; Riemenschneider, C.; Reemtsma, T.; Chefetz, B.; Hadar, Y., Transformation Pathways of the Recalcitrant Pharmaceutical Compound Carbamazepine by the White-Rot Fungus Pleurotus ostreatus: Effects of Growth Conditions. Environ. Sci. Technol. 2015, 49 (20), 12351-12362.

497 498 499

39. Lauchnor, E. G.; Semprini, L., Inhibition of phenol on the rates of ammonia oxidation by Nitrosomonas europaea grown under batch, continuous fed, and biofilm conditions. Water Res. 2013, 47 (13), 4692-4700.

500 501 502

40. Eichhorn, P.; Ferguson, P. L.; Pérez, S.; Aga, D. S., Application of ion trap-MS with H/D exchange and QqTOF-MS in the identification of microbial degradates of trimethoprim in nitrifying activated sludge. Anal. Chem. 2005, 77 (13), 4176-4184.

21



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503 504 505

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41. Jewell, K. S.; Castronovo, S.; Wick, A.; Falås, P.; Joss, A.; Ternes, T. A., New insights into the transformation of trimethoprim during biological wastewater treatment. Water Res. 2016, 88, 550-557.

506 507

22


Page 23 of 28

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

508


Table and figure legends

509 510

Table 1. Mass parameters for LC-MS/MS analysis.

511 512

Figure 1. (A) Nitrification performance of the enriched nitrifying sludge based on the cycle

513

studies at initial NH4+-N concentration of 20 mg L-1 and initial MLVSS concentration of 95±6.7

514

mg L-1 and (B) The calculated ammonia oxidation rate during the batch biodegradation

515

experiments with atenolol under different ammonium concentrations.

516 517

Figure 2. The effect of ammonium (NH4+-N) concentration on the degradation of atenolol (A)

518

and on the formation of its biotransformation product atenolol acid (B). C is the concentration of

519

atenolol or atenolol acid and C0 is the initial concentration of atenolol.

520 521

Figure 3. The concentration profiles of atenolol and atenolol acid normalized to the initial

522

concentration of atenolol during the time course in the experiments (A) with ammonia oxidation

523

(50 mg L-1 NH4+-N) and (B) without ammonia oxidation. C is the concentration of atenolol or

524

atenolol acid and C0 is the initial concentration of atenolol.

525 526

Figure 4. The relationships between ammonia oxidation rate and atenolol degradation rate (A);

527

between ammonia oxidation rate and atenolol acid formation rate (B).

528

23



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Page 24 of 28

529 530

Table 1. Mass parameters for LC-MS/MS analysis. Precursor ion

DP

Q1

CE (eV)

Q2

CE (eV)

(m/z)

(V)

(quantification)

/CXP (V)

(confirmation)

/CXP (V)

267.2

71

145.3

37/12

190.2

29/16

268.2

71

145.3

37/12

191.2

29/16

274.2

71

145.1

37/12

79.1

33/6

Compounds

Atenolol Atenolol acid Atenolol-d7 531

24


Page 25 of 28

532 24

70

(A)

25 mg L-1 of ammonium 50 mg L-1 of ammonium

(B) 60

Ammonia oxidation rate + -1 (mg NH4 -N gVSS h )

20

16 -1

(mg L )

NH+4-N/NO-2-N/NO-3-N concentration

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60


NH+4-N

12

2 3

NO -N NO -N

8

4

50 40 30 20 10 0

0 0

2

4

6

8

10

12

14

16

18

20

22

24

0

24

48

72

Time (h)

96

120

144

168

192

216

240

Time (h)

533 534 535

Figure 1. (A) Nitrification performance of the enriched nitrifying sludge based on the cycle

536

studies at initial NH4+-N concentration of 20 mg L-1 and initial MLVSS concentration of 95±6.7

537

mg L-1 and (B) The calculated ammonia oxidation rate during the batch biodegradation

538

experiments with atenolol under different ammonium concentrations.

539

25



540 120

120

Batch without ammonium Batch with 25 mg L-1 ammonium Batch with 50 mg L-1 ammonium

(A) 100

100

C/C0 (%)

80

60

60

40

40

20

20

0

0 0

541

Batch without ammonium Batch with 25 mg L-1 ammonium Batch with 50 mg L-1 ammonium

(B)

80

C/C0 (%)

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Page 26 of 28

24

48

72

96

120

144

168

192

216

240

0

24

48

72

96

Time (h)

120

144

168

192

216

240

Time (h)

542

Figure 2. The effect of ammonium (NH4+-N) concentration on the degradation of atenolol (A)

543

and on the formation of its biotransformation product atenolol acid (B). C is the concentration of

544

atenolol or atenolol acid and C0 is the initial concentration of atenolol.

545

26


Page 27 of 28

546 120

120

Atenolol Atenolol acid Mass balance

(A) 100

100

C/C0 (%)

80

60

60

40

40

20

20

0

0 0

547

Atenolol Atenolol acid Mass balance

(B)

80

C/C0 (%)

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60


24

48

72

96

120 144

168 192 216 240

0

24

48

72

Time (h)

96

120 144 168 192

216 240

Time (h)

548

Figure 3. The concentration profiles of atenolol and atenolol acid normalized to the initial

549

concentration of atenolol during the time course in the experiments (A) with ammonia oxidation

550

(50 mg L-1 NH4+-N) and (B) without ammonia oxidation. C is the concentration of atenolol or its

551

transformation product and C0 is the initial concentration of atenolol.

552

27



553 25 mg L-1 of ammonium 50 mg L-1 of ammonium

0.150

(A) Atenolol acid formation rate (µg atenolol acid g-1 h-1) VSS

0.225

Atenolol degradation rate (µg atenolol g-1VSS h-1)

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Page 28 of 28

0.180

0.135

y=0.0049x+0.013 R2=0.98

0.090

y=0.0034x-0.0029 R2=0.99

0.045

(B)

0.125 0.100 y=0.0031x-0.0014 R2=0.98

0.075

y=0.0019x+0.0016 R2=0.99

0.050 0.025 0.000

0.000 0

554

-1

25 mg L of ammonium 50 mg L-1 of ammonium

5

10

15

20

25

30

35

40

45

50

55

60

0

5

10

15

20

25

30

35

40

45

50

55

60

Ammonia oxidation rate (mg NH+4-N g-1VSS h-1)

Ammonia oxidation rate (mg NH+4-N g-1VSS h-1)

555 556

Figure 4. The relationships between ammonia oxidation rate and atenolol degradation rate (A);

557

between ammonia oxidation rate and atenolol acid formation rate (B).

558

28


Impact of Ammonium Availability on Atenolol ... - ACS Publications

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