Biodegradation of Chemicals in Unspiked Surface Waters

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Environmental Processes

Biodegradation of Chemicals in Unspiked Surface Waters Downstream of Wastewater Treatment Plants Zhe Li, and Michael S. McLachlan Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b05191 • Publication Date (Web): 23 Jan 2019 Downloaded from http://pubs.acs.org on January 26, 2019

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Environmental Science & Technology

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Biodegradation of Chemicals in Unspiked Surface Waters

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Downstream of Wastewater Treatment Plants

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Zhe Li* and Michael S. McLachlan

4

Department of Environmental Science and Analytical Chemistry (ACES), Stockholm University, S-10691

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Stockholm, Sweden

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TABLE OF CONTENTS ART

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8

ABSTRACT

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The OECD 309 guideline uses spiked incubation tests to provide data on biodegradation kinetics in

10

surface waters. However, potential limitations of spiking test chemicals into the studied water have not

11

been investigated. We conducted the OECD 309 test with unspiked surface water relying on chemical

12

residues present in the water. Parallel experiments were conducted with the same water spiked with 13

13

chemicals at higher concentrations (50 µg L-1). Six chemicals detected in both the spiked and unspiked

14

systems were biodegraded. For each chemical the concentration change over time differed between the

15

systems. Tramadol and venlafaxine showed constant concentrations in the spiked systems but

16

increasing concentrations in the unspiked systems. Atenolol and metoprolol showed first order

17

elimination with no lag in the unspiked systems, compared to a lag of 15-28 d followed by zero order

18

elimination kinetics in the spiked systems. Acesulfame was only slightly degraded (99%) in the spiked systems. Gabapentin displayed a

20

complex behavior where the features differed markedly between the spiked and unspiked systems. We

21

conclude that spiking can strongly influence biodegradation, reducing the environmental relevance of

22

test results. Under some conditions biodegradation can be measured in unspiked natural waters instead.

ACS Paragon Plus Environment

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INTRODUCTION

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The widespread presence of emerging organic contaminants in the aquatic environment has been well

25

documented during the last two decades.1-4 However, knowledge gaps still exist on their environmental

26

fate. Biodegradation is an important mechanism for removal of organic contaminants in natural

27

systems.5,6 For many compounds, the rate of biodegradation is a fundamental determinant of the

28

environmental fate of the compound.

29

Much of our understanding of aquatic biodegradation stems from laboratory biodegradation tests. These

30

tests have been developed with different applications in mind. One application of biodegradation testing

31

is to study biodegradability. Here the question is whether the substance of interest could conceivably

32

be degraded by microorganisms present in the water. This application is particularly interesting for

33

studying whether new chemicals are likely to be degraded in wastewater treatment plants (WWTPs).

34

One major research focus in recent years has been to investigate how the source and number of

35

microorganisms introduced into the test system influences the test results.7

36

A second application of biodegradation testing is to estimate the rate of biodegradation of chemicals in

37

the environment. The OECD 309 guideline (“Aerobic Mineralization in Surface Water”)8 is an

38

important test for providing kinetic data on biodegradation in surface waters for use in persistence

39

assessment and risk assessment. It measures biodegradation in aerobic natural waters that have been

40

spiked with test chemicals and incubated in the laboratory. One virtue of the OECD 309 simulation test

41

is that it employs natural waters. The degrading ability of microorganisms present in a specific water

42

body is measured, thus allowing the investigation of the system specific variability in biodegradation.

43

A constraint of the OECD 309 test is that it only assesses biodegradation in the water column, not in

44

the sediment or at the sediment-water interface. It can thus have limited relevance for assessing

45

biodegradation in aquatic systems in which biodegradation in sediment or at the sediment-water

46

interface dominates.9

47

A potential limitation of the OECD 309 test is that the test chemical is spiked into the surface water to

48

be studied, which could reduce the environmental relevance of the test system.10 There is evidence that

49

the biodegradation properties of the test system can be modified by adding chemicals. For instance, a

50

lag phase is often observed between addition of the test chemical and the beginning of measurable

51

degradation, suggesting that the microbial population is changing in response to the addition of the

52

chemical. There is also empirical evidence that biodegradation kinetics are influenced by the amount

53

of chemical added. For instance, the half-lives of 2,4-D and MCPP decreased by two orders of

54

magnitude and the lag time changed from 0 to 3-6 weeks when the spiked concentrations were increased

55

from 1 µg L-1 to 100 µg L-1.11 Many chemicals that are known to be degraded at high concentrations are

56

not degraded at environmentally relevant concentrations, even when capable degrading organisms are


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57

present.12 Moreover, it has been reported that kinetic rates measured at higher concentrations may

58

significantly differ from observations at much lower environmentally relevant concentrations.13

59

This study was designed to test whether it is possible to measure degradation kinetics in the OECD 309

60

test without spiking, relying instead on the chemical residues present in the surface water studied. In

61

addition, we aimed to compare the biodegradation kinetics from such an unspiked test with the

62

biodegradation kinetics obtained when test chemicals were spiked according to the OECD 309 guideline.

63

To this end, OECD 309 experiments were carried out using: a) lake water; b) lake water spiked with

64

test chemicals. A mixture of 13 test chemicals comprising a range of biodegradability was used for the

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spiked systems.

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MATERIALS AND METHODS

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Most of the experiments were conducted with water from Norra Bergundasjön (NB), a lake in southern

69

Sweden. This lake is the recipient for the Sundet WWTP, which serves the small city of Växjö. The

70

ratio of WWTP effluent discharge to freshwater inflow in this lake is ~1:4.

71

Biodegradation experiments were conducted for seven different test systems (see Table 1). Natural lake

72

water from NB was used for three test systems: an unspiked system to test if biodegradation can be

73

measured under unspiked conditions, a spiked system to allow comparison of biodegradation kinetics

74

under spiked and unspiked conditions, and a spiked abiotic system to correct for abiotic degradation

75

and sorption losses. Since the residence time of the water in NB was several months, we anticipated

76

that the concentrations of many of the test chemicals in NB would be low. Therefore, in order to increase

77

the likelihood of detecting the test chemicals in the unspiked system, we constructed three analogous

78

test systems using mixed lake water (referred to as mixture), which consisted of inflowing lake water

79

and WWTP effluent combined in proportion to their mixing ratio in the lake (4:1)). Finally, because the

80

wastewater effluent content of NB was high, we also constructed one test system using water from

81

Mälaren (MÄ), the third-largest freshwater lake in Sweden, to provide insight into test chemical

82

biodegradation in a less effluent-impacted surface water. In this case only a spiked system was

83

constructed, as we anticipated that the test chemicals would not be present in the MÄ surface water at

84

measurable concentrations.

85 86

Table 1. Conditions in the test systemsa system

water

treatment

unspiked natural NB

lake water from NB

none

pH

L-1

8.13 ± 0.19

DO (mg L-1) 8.26 ± 1.12

TSS (%) 3.4

spiked natural NB abiotic natural NB

lake water from NB lake water from NB

spiked at 50 µg spiked at 50 µg L-1; NaN3 at 0.1%

8.14 ± 0.12 8.52 ± 0.33

8.39 ± 0.54 9.04 ± 0.15

3.1 2.9

unspiked mixture NB

inflowing water and effluent (80:20, v/v)

none

8.20 ± 0.08

9.02 ± 0.32

3.1

spiked mixture NB


spiked at 50 µg L-1

8.24 ± 0.07

8.93 ± 0.18

3.0

abiotic mixture NB


8.53 ± 0.31

9.11 ± 0.29

3.2

spiked MÄ

lake water from MÄ

spiked at 50 µg L-1; NaN3 at 0.1% spiked at 50 µg L-1

8.02 ± 0.06

9.01 ± 0.29

0.6

87 88

a

89

experiment. Total suspended solids (TSS) was measured at the beginning of the experiment.

NB and MÄ stand for Lake Norra Bergundasjön and Lake Mälaren, respectively. The pH and dissolved oxygen

(DO) values represent mean values ± standard deviation calculated using the measurements from the whole

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Chemicals and Reagents

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The test chemicals (purity >98%) were purchased from Sigma-Aldrich (Steinheim, Germany) or

93

Toronto Research Chemicals Inc. (North York, Canada) and were stored under recommended

94

conditions until use. Details of these chemicals are provided as Table S1 in the Supporting Information

95

(SI). D- or

96

Chemicals Inc. and CDN Isotopes (Pointe-Claire Quebec, Canada) for use as internal standards. Stock

97

solutions of the standards were prepared in methanol and stored in amber CERTANÒ capillary bottles

98

in the dark at -20 °C. LC/MS-grade formic acid was purchased from Sigma-Aldrich. LC/MS-grade

99

acetonitrile and methanol were purchased from VWR (Stockholm, Sweden). Milli-Q water was

100

produced by using a Milli-Q Integral Water Purification System (Merck Millipore, Stockholm, Sweden).

101

A non-labeled standard solution containing all the 13 test chemicals and an internal standard solution

102

containing the 13 isotope-labeled standards at a concentration of 5 µg mL-1 were prepared in methanol

103

from the stock solutions; both standard solutions were stored in the dark at -20 °C until use.

104

Test Chemical Selection

105

Test chemical selection was based on several criteria: i) not volatile (i.e., log KAW 3:1 and >10:1, respectively (see Table S2). When the concentration of a chemical was 99.9

>99.9

N.D.b

>99.9

>99.9

3.6 c

4.9 c

caffeine

>99.9

>99.9

N.D.b

N.D.b

21.4

-2.2 c

-3.8 c

carbamazepine

1.2 c

0.9 c

-3.0 c

-11.5

-2.9 c

-2.6 c

-3.1 c

gabapentin

98.0

99.7

84.6

90.5

21.7

0.9 c

-1.2 c

hydrochlorothiazide

67.6

66.8

N.D.b

59.9

53.7

64.6

63.3

metoprolol

>99.9

89.9

N.D.b

>99.1

3.6 c

-0.9 c

-1.0 c

oxazepam

31.3

29.7

25.9

23.6

20.9

31.0

30.4

paracetamol

>99.9

>99.9

N.D.b

N.D.b

23.1

20.9

18.2

tramadol

2.2 c

4.8 c

-20.1

-9.5

-3.4 c

-1.6 c

-3.1 c

valsartan

>99.9

>99.9

N.D.b

N.D.b

2.7 c

-0.9 c

-3.0 c

venlafaxine

3.4 c

5.4 c

-59.9

-13.0

-2.3 c

-0.5 c

-1.6 c

chemical

spiked MÄ

230 231

a

232 233 234 235

water sample collected at the end of the incubation. In these cases, the > sign is used before the calculated

b

N.D. indicates that a chemical was not detected in any sample.

236

c

No significant dissipation occurred during the 60-day period (p >0.05 when testing the null hypothesis that the

237

slope of the linear regression of normalized concentration versus time is zero).

Dissipation (%) is the concentration of a chemical in the water at the end of the incubation relative to the

beginning of the experiment. LOQ was used for calculating dissipation when the chemical was not detected in the dissipation. A negative value indicates that the concentration of a chemical was higher at the end of the incubation than at the beginning.

238 239

Comparison of Chemical Dissipation Kinetics in Spiked and Unspiked NB Systems

240

The spiked and unspiked NB showed different dissipation kinetics for all chemicals for which

241

biodegradation was important (see Figure 1). The differences in dissipation kinetics varied between

242

chemicals. There was no common pattern that applied to all chemicals, but it was possible to group the

243

chemicals according to their behavior.


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Figure 1. Dissipation kinetics of all test chemicals in the spiked and unspiked NB systems: A) atenolol,

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B) metoprolol, C) gabapentin, D) carbamazepine, E) tramadol, F) venlafaxine, G) acesulfame, H)

247

hydrochlorothiazide, I) oxazepam, J) caffeine, K) paracetamol, and L) valsartan. Each data point is the

248

average of 3 measurements from each of two bottles. The scale on the y-axis is logarithmic and different

249

for each row. Error bars represent standard deviation (n=3).

250


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Table 3. Summary of dissipation half-life (t1/2, d) estimations of test chemicals in the test bottlesa chemical

spiked natural NB

spiked mixture NB

unspiked natural NB

unspiked mixture NB

spiked natural MÄ

acesulfame

3.0 (28)

2.7 (42)

48.0 (28)

77.7 (21)

--d

2.7 (28)

2.1 (42)

42.9 (21)

91.5 (21)

--d

N.Q.b (15)

N.Q.b (21)

N.D.c

9.3 (0)

N.Q.b (21)

N.Q.b (15)

N.Q.b (21)

N.D.c

10.7 (0)

N.Q.b (21)

1.9 (28)

2.7 (28)

N.D.c

N.D.c

78.1 (28)

2.1 (28)

4.0 (21)

N.D.c

N.D.c

81.0 (28)

--d

--d

--d

--d

--d

--d

--d

--d

--d

--d

N.Q.b (15)

N.Q.b (21)

N.Q.b (7)

N.Q.b (4)

--d

N.Q.b (15)

N.Q.b (21)

N.Q.b (7)

N.Q.b (4)

--d

37.0 (0)

37.7 (0)

N.D.c

38.6 (0)

77.8 (0)

35.4 (0)

36.4 (0)

N.D.c

43.1 (0)

72.5 (0)

N.Q.b (21)

N.Q.b (28)

N.D.c

16.0 (1)

--d

N.Q.b (21)

N.Q.b (28)

N.D.c

16.0 (0)

--d

116.2 (0)

119.3 (0)

130.3 (0)

127.7 (0)

170.4 (0)

114.0 (0)

124.1 (0)

123.1 (0)

131.2 (0)

159.9 (0)

0.9 (15)

2.5 (21)

N.D.c

N.D.c

112.5 (21)

0.9 (15)

2.6 (21)

N.D.c

N.D.c

147.2 (15)

--d

--d

--d

--d

--d

--d

--d

--d

--d

--d

1.7 (2)

0.7 (2)

N.D.c

N.D.c

--d

1.6 (2)

1.7 (2)

N.D.c

N.D.c

--d

--d

--d

--d

--d

--d

--d

--d

--d

--d

--d

atenolol caffeine carbamazepine gabapentin hydrochlorothiazide metoprolol oxazepam paracetamol tramadol valsartan venlafaxine

252

a

253

indicate the t1/2 calculated for individual replicate bottles; numbers in brackets indicate the duration of the lag

254

phase (d).

255

b

N.Q. indicates that t1/2 could not be estimated because first order elimination kinetics were not observed.

256

c

N.D. indicates that a chemical was not detected in any sample.

257

d

-- indicates that no significant dissipation occurred during the 60-day period (p >0.05 when testing the null

258

hypothesis that the slope of the linear regression of normalized concentration versus time is zero).

t1/2 (d) was estimated based on pseudo first-order kinetics excluding the lag phase; numbers in the two rows

259


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Atenolol and Metoprolol

261

Atenolol and metoprolol were removed in both the spiked and unspiked NB systems, but the shape of

262

the dissipation curve and the extent of removal differ (Figures 1A and 1B). Both chemicals show non-

263

first order dissipation in the spiked systems with a lag phase varying from 15 to 28 d. A lag phase of 21

264

d was also consistently observed for the dissipation of atenolol in the spiked MÄ bottles (Figure S3).

265

The lag suggests that biodegradation is governed by Monod population growth kinetics, whereby the

266

microorganisms capable of using the chemical as a growth substrate first have to reproduce before they

267

are sufficient in number to degrade a significant portion of the chemical.23 The OECD 309 guideline

268

recognizes the frequent occurrence of the lag phase and specifies that the half-life should be calculated

269

based on only the linear portion of the semi-logarithmic plot of concentrations versus time. However,

270

the plots never became linear, with the slopes continually increasing with time (Figures 1A and 1B).

271

Indeed, after the lag time the dissipation of atenolol and metoprolol was nearly zero order, which would

272

be consistent with Monod population growth kinetics with an excess supply of substrate23. It is

273

meaningless to calculate half-lives in such a situation,8 but it is clear that after the lag time dissipation

274

was much more rapid in the spiked systems than in the unspiked systems (Figures 1A and 1B).

275

In contrast, no lag phase was observed in the unspiked mixture NB bottles (dissipation of atenolol and

276

metoprolol could not be measured in the unspiked natural NB bottles). For both chemicals the

277

dissipation was first order either without any lag (atenolol) or with a short lag of 1 d (metoprolol). This

278

was consistent with expectations. In unspiked water where the environmental conditions and

279

contaminant levels are close to steady state, a lag phase would not be expected as the microbial

280

population would be expected to be adapted to the available substrates (including the chemicals of

281

interest).

282

In summary, the spiked tests gave a log lag phase and zero order kinetics while the unspiked test gave

283

no or little lag phase and first order kinetics. The spiked tests are clearly a poorer representation of the

284

situation in the real environment, as they cause a major adaptation of the microbial population before

285

the dissipation kinetics are measured. We conclude that for these chemicals the spiked bottle test did

286

not give a good measure of biodegradation kinetics in lake water.

287

Gabapentin

288

Remarkable differences in dissipation kinetics between the spiked and unspiked NB systems were

289

observed for gabapentin (Figure 1C). Similar to atenolol and metoprolol, gabapentin shows a long lag

290

time in the spiked systems followed by comparatively rapid first order dissipation. In contrast, in the

291

unspiked system gabapentin concentrations decreased rapidly after a shorter lag, reached a plateau after

292

1-3 weeks, and then increased slightly but statistically significantly. The plateau behavior could have

293

been caused by the formation of one or several persistent transformation products that blocked further

294

transformation of gabapentin, while the subsequent increase may have been due to an onset of


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295

gabapentin formation from certain precursor transformation products. Formation of parent organic

296

contaminants from both product-to-parent transformation24 and back-conversion of conjugated

297

metabolites via de-conjugation25,26 have been previously demonstrated. Such reactions are also

298

suggested to partially explain the negative removal efficiency results obtained for some emerging

299

organic contaminants from WWTPs.27-29 However, back-conversion of human metabolites does not

300

appear to be a viable explanation for gabapentin as it has been documented that gabapentin is not

301

metabolized at all in humans and is therefore excreted exclusively as the native substance via urine.30

302

Therefore, the formation of such transformation products would have to occur during wastewater

303

collection and treatment processes and/or later in the surface water. Biotransformation of gabapentin

304

has recently been studied and possible reaction pathways have been proposed.31 Several transformation

305

products were confirmed, at least two of which (gabapentin lactam and TP 213) can transform back to

306

the native chemical via amide hydrolysis and deacetylation under aerobic conditions in surface water.31

307

Another process that may have influenced the behavior of gabapentin was discovered by Gulde et al.32

308

They demonstrated that amines are sequestered into protozoa in activated sludge, and that this made

309

them unavailable for biodegradation. The gradual increase in gabapentin concentration during the latter

310

part of the incubation may have been a consequence of release from a declining protozoa population.

311

The plateau behavior followed by a concentration increase was not observed for the spiked mixture lake

312

water system, and in the spiked natural system it was only observed between day 42 and day 60. With

313

regard to the transformation product hypothesis above, this could be explained by the transformation

314

products (which were responsible for blocking further dissipation of gabapentin) having been persistent

315

only in the unspiked system, not in the spiked system. With regard to the protozoa sequestration effect,

316

the different plateau behavior could be explained by the protozoa sequestration saturating, resulting in

317

protozoa taking up a much smaller fraction of the gabapentin spiked into the bottle. These questions

318

could not be fully elucidated within the scope of this study. Nevertheless, the gabapentin case presents

319

another example of spiked tests giving a very different picture of biotransformation kinetics than

320

unspiked tests.

321

Neither the spiked nor the unspiked systems show removal of gabapentin from the start of the test. This

322

stands in contrast to atenolol and metoprolol, which were removed from the unspiked systems after no

323

or little lag phase. It is thus not clear if any of the complex elements of the dissipation behavior discussed

324

above can be attributed to the real environment. As with all laboratory tests there is some uncertainty

325

in the environmental relevance of the unspiked NB tests. For instance, there was a period of ~24 h

326

during which the water samples were stored before starting the tests, and it cannot be ruled out that the

327

system was disturbed such that a longer acclimatization phase was needed to restore the system

328

performance with respect to some chemicals. In situ measurements of the persistence of gabapentin in

329

the lake might help to resolve this uncertainty.


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Carbamazepine, Tramadol, and Venlafaxine

331

As stated above, the concentrations of carbamazepine, tramadol, and venlafaxine stayed unchanged

332

during the 60-day period in both the spiked natural and the spiked mixture NB systems (Tables 2 and 3

333

and Figures 1D-1F). The carbamazepine concentration was also constant in the unspiked natural NB

334

system, but in the unspiked mixture NB system the carbamazepine concentration increased by 11.5%

335

over the course of the incubation (Table 2). Tramadol and venlafaxine behaved similarly; their

336

concentrations increased in both the unspiked natural and unspiked mixture lake water systems.

337

Possible explanations include formation of the persistent parent chemicals from conjugates or

338

transformation products present in the water, or, for tramadol and venlafaxine, release of parent

339

chemical that had been sequestered by amine ion trapping into protozoa (in analogy to gabapentin). A

340

major well-documented metabolite of venlafaxine (O-desmethylvenlafaxine)22,33 was detected (Figure

341

2). Its concentration decreased over time in the unspiked systems, while it increased in the spiked

342

systems. It is also notable that the formation of venlafaxine in the unspiked systems as well as the

343

formation of its metabolite in the spiked systems was much more pronounced in natural lake water

344

compared to mixture lake water. This illustrates that there are limitations to using the mixture lake water

345

as a model for the natural lake water. Comparing the unspiked and spiked systems, the results illustrate

346

that the standard (spiked) OECD 309 test does not capture the formation of the target chemical from

347

previously formed precursors, and that for some chemicals this process can decisively influence

348

dissipation in surface water. However, the spiked tests do contribute valuable information, showing that

349

carbamazepine, tramadol, and venlafaxine are not degraded even after a 60-day incubation period.

350

351 352

Figure 2. Behavior of venlafaxine (A) and O-desmethylvenlafaxine (B) in the spiked and unspiked NB

353

systems. Concentrations at each time point were normalized to the measured initial values for each

354

chemical. Each data point is the average of 3 measurements from each of two bottles. The scale on the

355

y-axis is logarithmic and different for the two chemicals.


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356

Acesulfame

357

Acesulfame, a chemical that has been reported to be persistent and that is used frequently as tracer in

358

the aquatic environment,17,18,34,35 also behaved differently between the spiked and unspiked NB systems

359

(Figure 1G). It was persistent in all systems for the first 28 days, but thereafter the dissipation rates

360

diverged markedly. Acesulfame was completely removed (>99.9%) by day 42 from the spiked natural

361

NB system and by day 60 in the spiked mixture NB system with a half-life of 99.9% was found for paracetamol in the spiked

384

NB systems compared to the ~20% in the abiotic systems (Table 2). This indicates that paracetamol

385

underwent both biotic and abiotic dissipation processes with the former one being dominant. However,

386

none of the three chemicals was detected in the unspiked NB systems, hence comparison to the spiked

387

system was not possible.

388


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Environmental Relevance

390

We found little evidence to support the hypothesis that the biodegradation rate in surface water is closely

391

represented by the biodegradation rate measured in surface water spiked with the test chemical. Instead

392

we obtained evidence that spiking standardized bottle tests with chemicals at a concentration markedly

393

higher than their environmental concentrations can give different degradation rates than unspiked tests.

394

For all of our test chemicals that were detected and degraded via biotic reactions, considerable

395

discrepancies were observed in the dissipation kinetics between spiked and unspiked systems.

396

Explanations for the effects of spiking on biodegradation kinetics include: i) growth of the population

397

of active degrading organisms in the spiked system in response to the higher levels of substrate; ii)

398

adaptation of the microbial population to be able to utilize the chemical as a food source at high

399

chemical concentrations while at low concentrations chemicals are mainly transformed co-

400

metabolically by enzymes present to utilize other major carbon pools10; iii) concentration thresholds for

401

the induction of the enzymes required for metabolism; and iv) competition from other food sources at

402

low chemical concentrations10.

403

In a review of biodegradation test methods, Kowalczyk et al. concluded that current OECD tests are

404

not environmentally realistic.10 Our work shows that spiking the test chemicals is one factor

405

contributing to this lack of realism. Given that the OECD 309 experimental protocol acknowledges and

406

anticipates adaptation of the microbial population to the changed environment caused by the addition

407

of the test chemicals, this result is not surprising. Instead of measuring a degradation rate that represents

408

the degradation that was occurring in the natural water when sampled, the OECD 309 protocol is better

409

designed to establish whether the microbial population in the surface water could adapt to and degrade

410

the chemical if it was presented at high concentrations. In this respect the OECD 309 protocol can be

411

useful for assessing a spill scenario, while it is less relevant for assessing biodegradation under

412

conditions where the chemical is continually present.

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The spiking level may well influence the deviation of OECD 309 from environmental realism. The

414

spiking level in our experiments (50 µg L-1) is high compared to the concentrations of the studied

415

chemicals in surface waters (e.g., the lakes in this study). It is possible that lower spiking levels would

416

have given results that are more environmentally realistic. However, the diversity of effects observed

417

in this study indicate that it will be difficult to establish a spiking level that will assure environmentally

418

realistic results for broad groups of chemicals and diverse surface waters.

419

If the objective of the biodegradation test is a measure of chemical degradation in surface water where

420

the chemical is already present, conducting the experiment with unspiked systems is a sounder approach.

421

Our work shows that unspiked tests are possible under some conditions. However, they can only be

422

applied to existing chemicals and the concentrations in the studied water must be sufficiently high to

423

allow measurement of chemical dissipation. Sensitive analytical methods are thus a requirement for


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424

broader application of the unspiked test. Analytical sensitivity can be increased by using large volume

425

injection (employed in this study) and/or online solid phase extraction.

426

The large difference in biodegradation rates of most chemicals between MÄ and NB is an indication of

427

the broad range of biodegradation rates in surface water. For most chemicals it will not be possible to

428

apply the unspiked test across a broad range of environments due to the requirement for a relatively

429

high concentration of the chemical in the water. While the unspiked test can give a measure of

430

biodegradation in more contaminated surface waters, other tools will be needed to extrapolate this

431

knowledge to less contaminated surface water bodies. It also remains to be seen whether a more

432

environmentally relevant OECD 309 test can provide insight into biodegradation in aquatic systems in

433

which the degradation is occurring primarily in sediment or at the sediment-water interface.

434

In addition to these limitations, measuring biodegradation in unspiked systems also opens new

435

possibilities. When combined with non-target analysis strategies involving high resolution mass data

436

acquisition,37 one can acquire information on the degradation of many chemicals – even substances

437

currently unknown – from a single test. Creating a community database of MS data from unspiked

438

biodegradation tests could facilitate biodegradation assessment of new chemicals, accelerating

439

knowledge acquisition in this field.

440

ASSOCIATED CONTENT

441

Supporting Information

442

List of test chemicals and their physical-chemical properties, summary of LOD and LOQ of test

443

chemicals, trends of pH, DO, and conductivity in all test bottles, summary of detection and initial

444

concentrations of test chemicals in all unspiked systems, dissipation kinetics of all test chemicals in the

445

spiked abiotic control systems, dissipation of all test chemicals in the spiked MÄ system.

446

AUTHOR INFORMATION

447

Corresponding Author

448

*Phone: +46 8 674 7188. E-mail: [email protected].

449

Notes

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The authors declare no competing financial interest.


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Page 19 of 21


451

ACKNOWLEDGEMENTS

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The authors thank the Swedish Research Council Formas for funding this study and Anneli Andersson

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Chan and her colleagues at the Sundet Wastewater Treatment Plant for assistance in obtaining the

454

effluent water.

455

REFERENCES

456 457 458 459 460 461 462 463 464 465 466 467 468 469 470 471 472 473 474 475 476 477 478 479 480 481 482 483 484 485 486 487 488 489 490 491 492

(1) (2)

(3)

(4)

(5)

(6) (7)

(8) (9) (10)

(11)

(12)

(13)

Ternes, T. A. Occurrence of drugs in German sewage treatment plants and rivers. Water Res. 1998, 32 (11), 3245–3260. Hass, U.; Duennbier, U.; Massmann, G. Occurrence and distribution of psychoactive compounds and their metabolites in the urban water cycle of Berlin (Germany). Water Res. 2012, 46 (18), 6013–6022. Trautwein, C.; Berset, J.-D.; Wolschke, H.; Kümmerer, K. Occurrence of the antidiabetic drug Metformin and its ultimate transformation product Guanylurea in several compartments of the aquatic cycle. Environ. Int. 2014, 70, 203–212. Ruff, M.; Mueller, M. S.; Loos, M.; Singer, H. P. Quantitative target and systematic non-target analysis of polar organic micro-pollutants along the river Rhine using high-resolution massspectrometry – Identification of unknown sources and compounds. Water Res. 2015, 87, 145– 154. Writer, J. H.; Antweiler, R. C.; Ferrer, I.; Ryan, J. N.; Thurman, E. M. In-stream attenuation of neuro-active pharmaceuticals and their metabolites. Environ. Sci. Technol. 2013, 47 (17), 9781–9790. Radke, M.; Maier, M. P. Lessons learned from water/sediment-testing of pharmaceuticals. Water Res. 2014, 55, 63–73. Martin, T. J.; Goodhead, A. K.; Snape, J. R.; Davenport, R. J. Improving the ecological relevance of aquatic bacterial communities in biodegradability screening assessments. Sci. Total Environ. 2018, 627, 1552–1559. OECD Guidelines for the Testing of Chemicals 309. Organisation for Economic Co-operation and Development (OECD): Paris 2004, pp 1–21. Kunkel, U.; Radke, M. Fate of pharmaceuticals in rivers: Deriving a benchmark dataset at favorable attenuation conditions. Water Res. 2012, 46 (17), 5551–5565. Kowalczyk, A.; Martin, T. J.; Price, O. R.; Snape, J. R.; van Egmond, R. A.; Finnegan, C. J.; Schäfer, H.; Davenport, R. J.; Bending, G. D. Refinement of biodegradation tests methodologies and the proposed utility of new microbial ecology techniques. Ecotoxicol. Environ. Saf. 2015, 111, 9–22. Toräng, L.; Nyholm, N.; Albrechtsen, H.-J. Shifts in biodegradation kinetics of the herbicides MCPP and 2,4-D at low concentrations in aerobic aquifer materials. Environ. Sci. Technol. 2003, 37 (14), 3095–3103. Kolvenbach, B. A.; Helbling, D. E.; Kohler, H.-P. E.; Corvini, P. F.-X. Emerging chemicals and the evolution of biodegradation capacities and pathways in bacteria. Curr. Opin. Biotechnol. 2014, 27, 8–14. Helbling, D. E.; Hammes, F.; Egli, T.; Kohler, H.-P. E. Kinetics and yields of pesticide biodegradation at low substrate concentrations and under conditions restricting assimilable organic carbon. Appl. Environ. Microbiol. 2014, 80 (4), 1306–1313.


19


493 494 495 496 497 498 499 500 501 502 503 504 505 506 507 508 509 510 511 512 513 514 515 516 517 518 519 520 521 522 523 524 525 526 527 528 529 530 531 532 533 534 535 536 537 538 539

(14)

(15)

(16) (17)

(18)

(19)

(20)

(21)

(22) (23) (24)

(25)

(26)

(27)

(28)

(29)

Page 20 of 21

Method 1694: Pharmaceuticals and personal care products in water, soil, sediment, and biosolids by HPLC/MS/MS. U.S. Environmental Protection Agency: Washington, DC December 2007. Reemtsma, T.; Weiss, S.; Mueller, J.; Petrovic, M.; González, S.; Barceló, D.; Ventura, F.; Knepper, T. P. Polar pollutants entry into the water cycle by municipal wastewater: A European perspective. Environ. Sci. Technol. 2006, 40 (17), 5451–5458. Löffler, D.; Römbke, J.; Meller, M.; Ternes, T. A. Environmental fate of pharmaceuticals in water/sediment systems. Environ. Sci. Technol. 2005, 39 (14), 5209–5218. Li, Z.; Sobek, A.; Radke, M. Fate of pharmaceuticals and their transformation products in four small European rivers receiving treated wastewater. Environ. Sci. Technol. 2016, 50 (11), 5614–5621. Zou, H.; Radke, M.; Kierkegaard, A.; MacLeod, M.; McLachlan, M. S. Using chemical benchmarking to determine the persistence of chemicals in a Swedish lake. Environ. Sci. Technol. 2015, 49 (3), 1646–1653. Li, Z.; Sobek, A.; Radke, M. Flume experiments to investigate the environmental fate of pharmaceuticals and their transformation products in streams. Environ. Sci. Technol. 2015, 49 (10), 6009–6017. Schaper, J. L.; Posselt, M.; McCallum, J. L.; Banks, E. W.; Hoehne, A.; Meinikmann, K.; Shanafield, M. A.; Batelaan, O.; Lewandowski, J. Hyporheic exchange controls fate of trace organic compounds in an urban stream. Environ. Sci. Technol. 2018, 52 (21), 12285–12294. Rúa-Gómez, P. C.; Püttmann, W. Degradation of lidocaine, tramadol, venlafaxine and the metabolites O-desmethyltramadol and O-desmethylvenlafaxine in surface waters. Chemosphere 2013, 90 (6), 1952–1959. Ghattas, A.-K.; Fischer, F.; Wick, A.; Ternes, T. A. Anaerobic biodegradation of (emerging) organic contaminants in the aquatic environment. Water Res. 2017, 116, 268–295. Schwarzenbach, R. P.; Gschwend, P. M.; Imboden, D. M. Environmental Organic Chemistry, 2nd ed.; John Wiley & Sons: Hoboken NJ, 2003; pp 739–750. Qu, S.; Kolodziej, E. P.; Long, S. A.; Gloer, J. B.; Patterson, E. V.; Baltrusaitis, J.; Jones, G. D.; Benchetler, P. V.; Cole, E. A.; Kimbrough, K. C.; Tarnoff, M. D.; Cwiertny, D. M. Product-to-parent reversion of trenbolone: unrecognized risks for endocrine disruption. Sci. 2013, 342 (6156), 347–351. Douziech, M.; Conesa, I. R.; Benítez-López, A.; Franco, A.; Huijbregts, M.; van Zelm, R. Quantifying variability in removal efficiencies of chemicals in activated sludge wastewater treatment plants - a meta-analytical approach. Environ. Sci.: Processes Impacts 2018, 20 (1), 171–182. McCall, A.-K.; Bade, R.; Kinyua, J.; Lai, F. Y.; Thai, P. K.; Covaci, A.; Bijlsma, L.; van Nuijs, A. L. N.; Ort, C. Critical review on the stability of illicit drugs in sewers and wastewater samples. Water Res. 2016, 88, 933–947. Li, Z.; Undeman, E.; Papa, E.; McLachlan, M. S. High-throughput evaluation of organic contaminant removal efficiency in a wastewater treatment plant using direct injection UHPLCOrbitrap-MS/MS. Environ. Sci.: Processes Impacts 2018, 20, 561–571. Baalbaki, Z.; Sultana, T.; Metcalfe, C.; Yargeau, V. Estimating removals of contaminants of emerging concern from wastewater treatment plants: The critical role of wastewater hydrodynamics. Chemosphere 2017, 178, 439–448. Jelic, A.; Gros, M.; Ginebreda, A.; Cespedes-Sánchez, R.; Ventura, F.; Petrovic, M.; Barceló, D. Occurrence, partition and removal of pharmaceuticals in sewage water and sludge during wastewater treatment. Water Res. 2011, 45 (3), 1165–1176.


20

Page 21 of 21

540 541 542 543 544 545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560 561 562 563


(30) (31)

(32)

(33)

(34)

(35)

(36)

(37)

Gallimore, C.; Gidal, B. Gabapentin. In Atlas of Epilepsies; Springer London: London, 2010; pp 1753–1760. Henning, N.; Kunkel, U.; Wick, A.; Ternes, T. A. Biotransformation of gabapentin in surface water matrices under different redox conditions and the occurrence of one major TP in the aquatic environment. Water Res. 2018, 137, 290–300. Gulde, R.; Anliker, S.; Kohler, H.-P. E.; Fenner, K. Ion trapping of amines in protozoa: A novel removal mechanism for micropollutants in activated sludge. Environ. Sci. Technol. 2018, 52 (1), 52–60. 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. 2014, 72, 3–27. Buerge, I. J.; Buser, H.-R.; Kahle, M.; Müller, M. D.; Poiger, T. Ubiquitous occurrence of the artificial sweetener acesulfame in the aquatic environment: An ideal chemical marker of domestic wastewater in groundwater. Environ. Sci. Technol. 2009, 43 (12), 4381–4385. Liu, Y.; Blowes, D. W.; Groza, L.; Sabourin, M. J.; Ptacek, C. J. Acesulfame-K and pharmaceuticals as co-tracers of municipal wastewater in a receiving river. Environ. Sci.: Processes Impacts 2014, 16 (12), 2789–2795. Kahl, S.; Kleinsteuber, S.; Nivala, J.; van Afferden, M.; Reemtsma, T. Emerging biodegradation of the previously persistent artificial sweetener acesulfame in biological wastewater treatment. Environ. Sci. Technol. 2018, 52 (5), 2717–2725. Hollender, J.; Schymanski, E. L.; Singer, H. P.; Ferguson, P. L. Nontarget screening with high resolution mass spectrometry in the environment: Ready to go? Environ. Sci. Technol. 2017, 51 (20), 11505–11512.


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Biodegradation of Chemicals in Unspiked Surface Waters

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