original scientific paper
ISSN 1330-9862
https://doi.org/10.17113/ftb.58.04.20.6424
Influence of Diclofenac on Activated Sludge Bacterial Communities
in Fed-Batch Reactors
Barbara Kraigher* and
Ines Mandic-Mulec
University of Ljubljana, Biotechnical
Faculty, Department of Food
Science and Technology, Chair of
Microbiology, Večna pot 111, 1000
Ljubljana, Slovenia
Received: 24 June 2019
Accepted: 24 November 2020
SUMMARY
Research background. The occurrence and environmental toxicity of pharmaceuticals
have recently attracted increasing attention. Diclofenac is a highly consumed non-steroidal anti-inflammatory drug, which is often detected in wastewaters, but investigations of
its influence on bacteria are scarce.
Experimental approach. We investigated the influence of this pharmaceutical on bacterial community in activated sludge exposed to increasing concentrations of diclofenac in
fed-batch reactors over 41 days. Nitrification activity of the activated sludge was measured
and changes in bacterial community structure were followed using culture-independent
molecular method (terminal restriction fragment length polymorphism, T-RFLP) and by
the cultivation approach.
Results and conclusions. Nitrification activity was not detectably influenced by the addition of diclofenac, while the main change of the bacterial community structure was detected only at the end of incubation (after 41 days) when diclofenac was added to artificial
wastewater as the only carbon source. Changes in community composition due to enrichment were observed using cultivation approach. However, taxonomic affiliation of isolates
did not match taxons identified by T-RFLP community profiling. Isolates obtained from
activated sludge used as inoculum belonged to five genera: Comamonas, Arthrobacter,
Acinetobacter, Citrobacter and Aeromonas, known for their potential to degrade aromatic
compounds. However, only Pseudomonas species were isolated after the last enrichment
step on minimal agar plates with diclofenac added as the sole carbon source.
Novelty and scientific contribution. Our results suggest that the selected recalcitrant
and commonly detected pharmaceutical does not strongly influence the sensitive and
important nitrification process of wastewater treatment. Moreover, the isolated strains
obtained after enrichment procedure that were able to grow on minimal agar plates with
diclofenac added as the only carbon source could serve as potential model bacteria to
study bacterial diclofenac degradation.
Key words: activated sludge, pharmaceuticals, diclofenac, bacterial community T-RFLP,
Pseudomonas
INTRODUCTION
*Corresponding author:
Phone: +38613203410
Fax: +38612573390
E-mail: barbara.kraigher@bf.uni-lj.si
402
Pharmaceuticals, generally regarded as being extremely advantageous for increasing
human life expectancy, on the other hand represent a serious burden for the environment.
Worldwide consumption rates of hundreds of tonnes of pharmaceuticals have a potential
to cause undesirable ecological and human health effects (1–3). The pharmaceutical industry has increased enormously in recent years and currently represents one of the most
profitable branches of industry. Consequently, substantial amounts of pharmaceuticals
can reach and impact the environment, either through direct discharge of pharmaceuticals
or due to inefficient elimination in wastewater treatment plants (WWTPs). Several studies
have demonstrated that some pharmaceuticals are efficiently eliminated by WWTPs (ibuprofen, naproxen, ketoprofen), while others are resistant to biodegradation (diclofenac,
clofibric acid) (4–7). Although many pharmaceuticals and their metabolites have been
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Food Technol. Biotechnol. 58 (4) 402–410 (2020)
detected in aquatic environments and even in drinking water
samples at concentrations ranging from ng/L up to hundreds
of μg/L (4,8–10), there still remains a limited understanding of
the extent and magnitude of their ecological effects.
Diclofenac is a non-steroidal anti-inflammatory drug
(NSAID) prescribed in daily doses of 75–150 mg to reduce inflammation. It acts as an analgesic to reduce pain in patients
with arthritis or acute injury and can even be used to reduce
menstrual pain (11). Over the counter use of diclofenac is approved in some countries for minor pains and fever associated with common infections, and it has been estimated that
940 tonnes are consumed annually worldwide (11). Chemically, diclofenac is 2-[(2,6-dichlorophenyl)amino]-benzeneacetic
acid. One benzene ring is chlorinated, which contributes to
the recalcitrant nature of this pharmaceutical for biological
degradation. The highly variable and relatively low biodegradation of diclofenac in WWTPs has been reported, mainly in the range of 21–40 % (11). In the aquatic environment,
diclofenac is one of the most frequently detected pharmaceuticals (12). It has been detected in drinking water (12) and
has caused the death of millions of vultures in Southeast Asia
through its use in veterinary medicine (13,14). In the liver, kidneys and gills of rainbow trout, cytopathology occurred already at concentrations of 1 μg/L (15). Studies of diclofenac
influence on microbial communities are scarce. Microorganisms in lotic biofilms were inhibited at diclofenac concentrations of 100 μg/L (16). At very high concentrations (50–100
mg/L), it was shown to inhibit bacterial growth in most of the
tested Gram-positive and Gram-negative strains by inhibiting
DNA synthesis (17).
In our previous studies (18,19), we have shown that a mixture of selected commonly used pharmaceuticals including
diclofenac caused shifts in the community structure of activated sludge in pilot bioreactors. The aim of this study is to
evaluate the effect of diclofenac, the most recalcitrant drug
in the pharmaceutical mixture (5), on the nitrification activity
and the community structure of activated sludge in fed-batch
reactors exposed to increasing concentration of diclofenac.
Additionally, we used the enrichment culture approach to isolate bacteria that are capable of growing on minimal agar
plates with diclofenac added as the only carbon source.
MATERIALS AND METHODS
Experimental setup and sampling
Batch cultures of activated sludge were prepared in 500-mL Erlenmeyer flasks containing 200 mL of artificial wastewater, which was prepared by dissolving a mineral nutrient
composition in tap water, simulating the composition of real
municipal wastewater (5). The following chemicals were used:
yeast extract (130 mg/L), casein peptone (130 mg/L), meat extract (130 mg/L; all Biolife Italiana, Milano, Italy), CH3COONH4
(317 mg/L), NH4Cl (40 mg/L; both Kemika, Zagreb, Croatia), K 2HPO4 (24 mg/L), KH2PO4 (8 mg/L), CaCO3 (100 mg/L),
MgCO3 (100 mg/L), NaCl (40 mg/L; all Sigma-Aldrich, Merck,
Gillingham, UK) and FeSO4·7H2O (5 mg/L; Merck, Darmstadt,
Germany). Each flask (fed-batch reactor) was inoculated with
20 mL of activated sludge obtained from the R200 reactor
described by Kraigher et al. (18), which was operated for 18
months with artificial wastewater containing 200 μg/L of
each of the following pharmaceuticals: diclofenac, ibuprofen,
naproxen, ketoprofen and clofibric acid. This sludge was selected because it contained a biomass already adapted to the
selected concentration of pharmaceuticals and it was expected that potential changes in community structure and enrichment of bacteria capable of degrading diclofenac would
be detected sooner. Fed-batch reactors were incubated at
220 rpm and 28 °C in the darkness. Activated sludge was exposed to either 200 μg/L diclofenac (Sigma-Aldrich, Merck)
throughout the entire experiment, or to 5 mg/L diclofenac
(with increasing concentrations from 1 to 5 mg/L in the first
10 days of incubation). One fed-batch reactor was incubated
without the addition of pharmaceuticals. The fed-batch reactors were operated in the following way throughout the experiment: the hydraulic retention time of two days was maintained by settling down the sludge every two days (three
days over the weekends), decanting the medium, and adding fresh artificial wastewater with (or without) diclofenac.
Each time the samples of influents (artificial wastewater) and
effluents (decanted medium) were taken and kept frozen until N-NH4+ and N-(NO2–+NO3–) concentrations were analysed
by continuous flow analyser (FlowSys Alliance Instruments,
Salzburg, Austria). After 20 days of incubation, the artificial
wastewater without organic carbon (i.e. without casein peptone, CH3COONH4, meat and yeast extract) was used to feed
the fed-batch reactors which were incubated for additional
three weeks. Activated sludge for bacterial community analyses was sampled (1 mL) at indicated time intervals and stored
at –80 °C until DNA extraction.
Bacterial community analyses
Terminal restriction fragment length polymorphism
(T-RFLP) analysis was used to compare bacterial community
structures. DNA was extracted from 1-mL subsamples (two
independent DNA isolations were obtained for each sample) using the UltraClean soil DNA isolation kit (MoBio Solano
Beach, CA, USA) according to the manufacturer’s instructions.
Isolated total community DNA (concentration of 50–100 ng/
μL) was checked on a 1 % agarose gel and compared to the
Gene Ruler DNA Ladder Mix (Thermo Fisher Scientific Baltics, UAB, Vilnius, Lithuania) to estimate the size and concentration. Purified DNA was then used as a template for polymerase chain reaction (PCR) with 16S rRNA gene primers
27f labelled with 6-FAM (6-carboxyfluorescein) at the 5’ end
and 927r (20) for the total community analyses, following the
protocols described by Kraigher et al. (18). Restriction with restriction enzyme MspI (Thermo Fisher Scientific Baltics, UAB),
ethanol precipitation and T-RFLP profiling were performed.
Profiles were generated using ABI PRISM 310 DNA sequencer and GeneScan analysis software (21). Terminal restriction
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B. KRAIGHER and I. MANDIC-MULEC: Influence of Diclofenac on Sludge Bacterial Communities
fragments (T-RFs) with peak heights of less than 50 fluorescence units and T-RFs that were less than 50 bp long were
excluded from the analyses. T-RFLP profiles of duplicate DNA
isolations per sample were analysed using the BioNumerics
program v. 6.6 (22). Normalized intensity values and the positions of the detected bands were used for cluster analysis,
while peaks representing less than 1 % of the total fluorescence of all peaks in a sample were excluded from the analysis. Pearson’s correlation coefficient, which considers both
fragment length and peak height, was used to calculate similarity coefficients. A dendrogram was then constructed using
the unweighted pair-group method with arithmetic means
(UPGMA) algorithm.
Isolation of bacteria capable of growing
on minimal plates with diclofenac
Bacteria were isolated from the activated sludge that was
used as inoculum for the incubation experiment and from
the fed-batch cultures at the end of the experiment (after
41 days of incubation with diclofenac, with the last 20 days
without the addition of any other external carbon source except for diclofenac). Mineral medium (MM) agar plates supplied with 20 mg/L of diclofenac as the only carbon and
energy source were used for isolation. MM agar plates consisted of (in g/L): 3.5 Na2HPO4·2H2O (Sigma-Aldrich, Merck),
1 KH2PO4 (Sigma-Aldrich, Merck), 0.5 (NH4)2SO4·2H2O (Merck),
0.1 MgCl2·6H2O (Honeywell-Fluka, Basel, Switzerland), 0.05
Ca(NO3)2·4H2O (Honeywell-Fluka), 15 highly purified agar (Biolife Italiana) at pH=7.25 and finally 1 mL trace element solution SL-4 was added. The trace element solution SL-4 consisted of (in g/L): 0.5 EDTA (Sigma-Aldrich, Merck), 0.2 FeSO4·7H2O
(Merck), 0.01 ZnSO4·7H2O (Merck), 0.003 MnCl2·4H2O (Merck),
0.03 H3BO3 (Merck), 0.02 g CoCl2·6H2O (Honeywell-Fluka),
0.001 g CuCl2·2H2O (Sigma-Aldrich, Merck), 0.002 NiCl2·6H2O
(Sigma-Aldrich, Merck) and 0.003 Na2MoO4·2H2O (Kemika).
The activated sludge was centrifuged, washed with MM medium and then different dilutions were plated on MM plates
with diclofenac and incubated in darkness at 28 °C. Distinct
colonies were reinoculated on R2A plates (R-2A agar; Fluka Analytical, Sigma-Aldrich Chemie GmbH, Merck) several
times until pure cultures were obtained, and then verified for
growth on MM plates with diclofenac as a sole carbon source.
Finally, purified colonies were grown overnight in yeast extract broth consisting of (in g/L): 3 yeast extract and 6 peptocomplex (Biolife Italiana), then frozen in glycerol (15 %) and
stored at –80 °C.
Identification of the isolated bacteria
The isolated bacteria were grown in yeast extract broth
(Biolife Italiana), overnight at 28 °C and 220 rpm in darkness.
The overnight cultures were then boiled for 10 min and 3 μL
of boiled cells were used as a template to amplify the 16S
rRNA genes using primers 27f and 1406R (23) and PCR conditions as described above for T-RFLP. PCR reactions were
404
purified with QuickClean 5M PCR purification kit (GenScript
USA Inc., Piscataway, NJ, USA) and sequenced by Macrogen
Inc. (Seoul, Korea). Sequences of 700–800 bp were obtained.
Some isolates were selected for sequencing with both forward and reverse primers (which gave approx. 300 bp of overlapping sequence) to get more reliable information about the
identity of the isolates. The partial 16S rRNA gene sequences
were identified by Basic Local Alignment Search Tool (BLAST;
(24)) searches at NCBI and deposited in the GenBank database
under accession numbers JF928537 to JF928555. GenBank sequences, which were identified by BLAST searches to be most
similar to the isolate sequences, were downloaded and included in the phylogenetic tree reconstruction using neighbour-joining method with 500 bootstrap replicates and the
Maximum Composite Likelihood evolutionary model within
the MEGA v. 4 (25).
RESULTS AND DISCUSSION
Nitrification activity during the experiment
Studies of xenobiotics in activated sludge generally focus on their degradation efficiency. However, it is also very
important to investigate the influence of xenobiotics on microbial activities responsible for the key functions of WWTPs,
such as nitrogen removal. In the previous study (19), a higher concentration of nitrate was detected in the reactors containing 50 μg/L of the pharmaceutical mix than in the reactors with lower concentration of pharmaceutical mix (5 or 0
μg/L), while no significant influence of pharmaceuticals on
ammonium removal was observed. In this study, nitrification
activity of the activated sludge in the fed-batch reactors exposed to diclofenac was monitored every two to three days
after settling down the sludge by measuring ammonium, nitrite and nitrate concentrations in the medium (Fig. 1). Nitrification performance as reflected by ammonium removal
was high in all batch reactors (nearly 100 %), irrespective of
the diclofenac concentration (Fig. 1a). A short adaptation period was observed during the first 4–6 days when the seeding sludge was transferred from the pilot reactor to the fed-batch reactors without diclofenac or its lower concentration
(200 μg/L). This lag period was not observed in the fed-batch
reactors with a high concentration of diclofenac (1 mg/L at
the beginning). However, irrespective of diclofenac concentration, almost no ammonium (mostly less than 0.5 mg/L) remained in the fed-batch reactors after six days of incubation.
A longer adaptation period for nitrite oxidation was observed
as its accumulation was detected for 13–15 days (Fig. 1b). After this period, nitrite almost totally disappeared, while nitrate concentrations increased. After the first 20 days, when
all carbon sources except diclofenac were excluded and the
influent ammonium concentration was lower (it dropped
from approx. 65 to approx. 9.5 mg/L when all carbon sources were eliminated), the concentration of nitrate fell rapidly
and was relatively constant in all fed-batch reactors until the
end of the experiment (Fig. 1c). In the previous study (19), the
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Food Technol. Biotechnol. 58 (4) 402–410 (2020)
a)
Influent
γ(ammonium)/(mg/L)
70
0
60
Dc200
50
Dc
40
30
20
10
0
0
4
8
12
16
20
24
28
32
36
40
t/day
b)
30
Influent
0
γ(nitrite)/(mg/L)
25
Dc200
20
Dc
15
10
5
0
0
4
8
12
16
20
24
28
32
36
first four days of experiment (samples designated t0, 2d, 4d),
clustered separately from the samples after 20 days (samples
designated 20d) and after 41 days (41d) of incubation, indicating shifts in bacterial communities over time. This could
partly be attributed to the adaptation to the altered conditions when transferring the sludge from the pilot reactors
to the fed-batch reactors (communities diverged to approx.
65 % similarity). The highest divergence (only approx. 40 %
similarity) was observed when all carbon sources were excluded from the artificial wastewater and diclofenac served
as the only carbon source (Fig. 2). Also, the community profile of the sludge incubated without diclofenac diverged
most from the other two communities at the end of the experiment. The similarity among the replicate samples was at
least 80 %, while similarity between fed-batch reactors with
or without diclofenac at the end of the experiment was approx. 55 % (Fig. 2).
40
t/day
c)
γ(nitrate)/(mg/L)
0
100
80
60
Similarity/%
Influent
100
Dc 41d A
80
Dc200
Dc 41d B
60
Dc
Dc200 41d A
Dc200 41d B
40
0 41d A
20
0 41d B
0 20d A
0
0
4
8
12
16
20
24
28
32
36
40
t/day
Fig. 1. The concentrations of: a) ammonium, b) nitrite and c) nitrate in
the small fed-batch reactors during the experiment. Vertical dashed
lines indicate the time point where all carbon sources except diclofenac were excluded from the influent artificial wastewater. Dc=
flask with increasing diclofenac concentration up to 5 mg/L, Dc200=
flask with γ(diclofenac)=200 μg/L, 0=flask without diclofenac, influent=artificial wastewater used to feed the reactors
0 20d B
Dc 20d A
Dc 20d B
Dc200 20d A
Dc200 20d B
Dc 4d A
0 2d A
Dc 2d A
t0 A
t0 B
concentration of nitrate in the effluent of the reactors was
very low (except in the reactors with 50 μg/L of pharmaceuticals). This was attributed to the denitrification in the reactors under the microaerophilic conditions that were probably present in the activated sludge flocs. In contrast, in this
experiment the reactors were thoroughly shaken and flocs
could not develop. However, despite different conditions in
the pilot reactors and fed-batch reactors used in this experiment, ammonium removal was high in both systems and inhibition of nitrification process by diclofenac was not detected.
Bacterial community shifts during the experiment
The bacterial community structure of activated sludge in
the fed-batch reactors was evaluated by T-RFLP at the beginning of the experiment, after 20 days of incubation with diclofenac (0.2 or 5 mg/L) and at the end of the experiment (41
days), when the sludge was incubated for the last 21 days with
diclofenac as the only carbon source. Comparison of T-RFLP
profiles of bacterial communities is shown as a dendrogram
in Fig. 2. Community profiles representing inoculum and the
Fig. 2. Comparison of the activated sludge bacterial community profiles treated with different concentrations of diclofenac at three different times of incubation. The dendrogram of the terminal restriction fragment length polymorphism (T-RFLP) profiles was generated
based on Pearson’s correlation using the unweighted pair-group
method with arithmetic means (UPGMA) method. 0=reactor without
diclofenac, Dc200=reactor with γ(diclofenac)=200 μg/L, Dc=reactor
γ
with increasing diclofenac concentration up to 5 mg/L, t0=activated
sludge used for seeding the flasks, A and B=DNA isolations from duplicate samples, 2d, 4d, 20d, 41d=samples taken after 2, 4, 20 and 41
days of the incubation
Fig. 3 shows the T-RFLP profiles of some representative
samples. Major differences could be observed in the profiles
at the end of incubation, where at least two new peaks are
evident. One of the two T-RF peaks was present in all samples
at the end of the experiment (on the right end of the profiles,
the length was not estimated since it was already above the
range of our standard, which was 500 bp) and could be attributed to the absence of carbon sources during the last 21
days of incubation. We affiliated the most distinctive peak,
which clearly appeared only in the fed-batch reactors with
diclofenac and not in the control (Fig. 3), with T-RF of 292 bp.
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B. KRAIGHER and I. MANDIC-MULEC: Influence of Diclofenac on Sludge Bacterial Communities
Fig. 3. Terminal restriction fragment length polymorphism (T-RFLP) profiles of the bacterial communities in the flasks at different times of incubation. The most distinctive change in the T-RF of 292 bp (indicated by an arrow) occurred in the flasks treated with diclofenac without any
other carbon source. The other T-RF indicated occurred in all flasks at the end of the experiment. t0=activated sludge used for seeding the flasks,
0=reactor without diclofenac, Dc200=reactor with γ(diclofenac)=200 μg/L, Dc=reactor with increasing diclofenac concentration up to 5 mg/L,
2d, 20d, 41d=samples taken after 2, 20 and 41 days of incubation experiment
This peak was identified by in silico T-RFLP analyses of the
clone library constructed
previouslyμ (18), considering the T-RF
γ
drifts of in silico predicted and the experimental T-RF lengths
of approx. 4 bp (19). Based on our previous analyses (19) the
292 bp T-RF could be affiliated with Acidobacteria phylum.
Representatives of this phylum are very difficult to isolate
(only a few representatives have been isolated so far) and
are therefore not well characterized metabolically, although
molecular analyses indicate their widespread presence across
variety of environments (26,27).
Our results indicated that there was no significant influence of dicofenac on the bacterial community structure
in activated sludge during the first 20 days of the experiment, when additional carbon was still present, even if its
concentration was 5 mg/L. This is not surprising as the community has been previously adapted to a variety of pharmaceuticals. Moreover, consistently with Zhang et al. (11), who
reported relatively low degradation efficiency for diclofenac
in WWTPs, bacteria will preferentially degrade more easily
degradable carbon source. However, a detectable change in
bacterial community structure occurred when diclofenac was
provided as the only carbon source and other carbon sources
were removed. This change indicates that one or more microbial types became enriched in the setting and that adapted
sludge has a potential for growth on diclofenac and consequently its removal. However, we cannot exclude the possibility that this shift might be the result of enrichment of certain
diclofenac-tolerant community members that grow on dead
biomass. Therefore, further studies are needed to characterize the species responsible for the changed pattern.
Isolation of bacteria capable of growth on minimal
agar plates with diclofenac as a sole carbon source
As we observed a change in microbial community structure due to diclofenac, we aimed to isolate the bacteria capable of growing on minimal medium containing diclofenac
as the only carbon source. Bacteria were isolated from the
original bioreactor sludge used for inoculating the fed-batch
reactors and from the sludge exposed to higher concentrations of diclofenac after the 41-day experiment. From the first
sample, representatives of five genera were isolated, while
406
after incubation in the absence of other sources of carbon
during the last 21 days of experiment, only two different
Pseudomonas sp. were isolated. In Fig. 4, a phylogenetic tree
indicates the affiliation of the isolates obtained in this study
with the related strains found by BLAST searches of the 16S
rRNA genes. Eleven isolates from the seeding sludge designated Dc1-Dc11 belonged to Gammaproteobacteria, Betaproteobacteria and Actinobacteria. From the activated sludge in
fed-batch reactors, after enrichment experiment we isolated
eight strains (Dc21–Dc28) which were taxonomically less diverse and belonged to Pseudomonas sp. (Gammaproteobacteria). Changes in the community structure before and after the
enrichment with diclofenac as the only carbon source were
detected by culture-independent molecular methods as well
as with the selective isolation of bacteria capable of growing
on diclofenac. Affiliations and identities of partial 16S rRNA
genes of the isolates with the related strains are shown in Table 1. Five species isolated from the seeding sludge showed
99–100 % sequence identity to related strains in the GenBank,
which harbour representatives previously indicated for their
potential to degrade aromatic compounds. Majority (7) of the
obtained isolates were relatives of Comamonas testosteroni,
which is able to mineralize complex and xenobiotic aromatic compounds as the sole carbon and energy source (28,29).
Boon et al. (30) demonstrated that wastewater reactors augmented by C. testosteroni I2 gfp showed accelerated degradation of toxic chlorinated aromatic compound 3-chloroaniline.
17
This strain also protected the nitrifying bacterial community,
allowing for its faster recovery from the toxic shocks. Acinetobacter sp. and Arthrobacter sp. HY2 degraded p-nitrophenol
(31,32), while Citrobacter strains isolated from various effluent
treatment plants degraded a range of aromatic compounds
and biotransformed mono- and dichlorophenols (33). Species
from the genus Aeromonas are commonly reported in clinical studies, but they also degrade aromatic compounds (34).
Although the same method of isolation was used for all
isolates in this work, the isolates from the seeding sludge
were more diverse than those exposed to diclofenac without other carbon sources. The seeding sludge from the reactor was exposed for 18 months to diclofenac (in a mixture of
five different pharmaceuticals at concentration of 200 µg/L)
with many other carbon sources always present in artificial
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Food Technol. Biotechnol. 58 (4) 402–410 (2020)
49
88
0.01
Dc23 JF928551
P. putida F1 NC 009512
Dc24 JF928552
Dc25 JF928553
Dc29 JF928555
100
P. putida KT2440 NC 002947
Dc27 JF928547
Pseudomonas sp.
P. putida PNP1 EU874452
Dc26 JF928554
92
P. putida PL2 FJ204928
Dc21 JF928550
100
P. fluorescens PC17 AY538263
Gammaproteobacteria
Dc28 JF928548
100
Pseudomonas sp. R20808 AJ786776
P. aeruginosa LESB58 NC 011770
100 Dc7 JF928542
C. gillenii AF025367
A. hydrophila 12H341 AB473001
98
100 Dc3 JF928538
74 Aeromonas media 480a FJ168773
Acinetobacter sp. bk2 HQ538658
Dc10 JF928545
100
Acinetobacter sp. E9 FJ392124
Comamonas sp. PD3 EF373535
Dc1 JF928549
C. testosteroni I2gfp EF421407
100 Dc2 JF928537
Dc6 JF928541
Betaproteobacteria
67 Dc4 JF928539
Dc9 JF928544
Dc8 JF928543
12 Dc11 JF928546
Arthrobacter sp. TD3 EF468655
A. protophormiae SSCT48 AB210984
100
Actinobacteria
59 Dc5 JF928540
85 Arthrobacter sp. R4 FN392695
Fig. 4. Phylogenetic tree based on partial 16S rRNA gene sequences showing the affiliations of the isolates obtained in this study (printed bold)
to similar known species found by Basic Local Alignment Search Tool (BLAST) searches in the GenBank database (22). The percentage of replicate
trees in which the associated taxa clustered together in the bootstrap test (500 replicates) is shown at the nodes
Table 1. Identification and similarity of the isolates with the known species as found by Basic Local Alignment Search Tool (BLAST) searches in
the GenBank (22)
Sludge after
enrichment
Seeding activated sludge
Strains with sequences similar to the isolates
(BLAST search)
Comamonas testosteroni I2gfp (EF421407)
Similarity/%
Strain information found in the GenBank
1299/1299 (100)
Comamonas sp. PD3 (EF373535)
1297/1299 (99.9)
Aeromonas media 480a (FJ168773)
1309/1311 (99.9)
Arthrobacter sp. R4 (FN392695)
Arthrobacter sp. TD3 (EF468655)
A. protophormiae SSCT48 (AB210984)
1272/1272 (100)
1276/1286 (99.2)
1274/1285 (99.1)
Citrobacter gillenii (AF025367)
1304/1305 (99.9)
Acinetobacter sp. E-9 (FJ392124)
1261/1263 (99.8)
Acinetobacter sp. bk_2 (HQ538658)
P. putida PNP-1 (EU874452)
P. putida PL2 (FJ204928)
1299/1301 (99.9)
1299/1299 (100)
1298/1299 (99.9)
P. putida F1 (NC009512)
1299/1299 (100)
P. putida KT2440 (NC002947)
P. aeruginosa LESB58 (NC011770)
1298/1299 (99.9)
1256/1308 (96.0)
wastewater, which enabled the growth of co-metabolizing bacteria, while in the fed-batch reactors, enrichment of
bacteria able to utilize diclofenac as the sole carbon source
was favoured. Seven out of eight Pseudomonas isolates were
3-chloroaniline-degrading Comamonas testosteroni strain,
I2gfp, isolated from wastewater
phenol-degrading bacteria
analysis of 16S rRNA gene mutations in a subset of
Aeromonas strains
microcystin-degrading bacteria isolated from water
2,4,6-trinitrotoluene-degrading bacteria
bacteria isolated from sewage sludge compost
Citrobacter phylogeny by 16S rRNA gene sequence
comparison
phenol-degrading bacteria isolated from leaf microbial
communities
representative strains isolated from bulking activated sludge
p-nitrophenol-degrading bacteria isolated from soil
strain capable of biodegrading polycyclic aromatic
hydrocarbons isolated from polluted soil
completely sequenced Pseudomonas putida F1, isolated from
a polluted creek in Urbana, IL, USA
complete genome sequence of Pseudomonas putida KT2440
the Liverpool epidemic strain of Pseudomonas aeruginosa
affiliated with Pseudomonas putida (Fig. 4), which is a metabolically versatile bacterium closely related to P. aeruginosa
and important for bioremediation of sites contaminated with
multiple aromatic hydrocarbons. The partial 16S rRNA gene
October-December 2020 | Vol. 58 | No. 4 407
B. KRAIGHER and I. MANDIC-MULEC: Influence of Diclofenac on Sludge Bacterial Communities
sequence of our isolate was 100 % similar to that of P. putida
F1, which is one of the most well-studied aromatic hydrocarbon-degrading bacterial strains (35).
Importance of the isolated bacteria
Many bacteria which are able to degrade different pollutants are being isolated from various polluted environments.
Our results suggest that the same species or genera previously characterized to be metabolically versatile and able
to degrade diverse aromatic compounds, emerging in the
environment, might also be able to degrade diclofenac. In
comparison with ibuprofen, the diclofenac phenol ring is
halogenated, which contributes to the recalcitrance of this
drug. However, when we performed similar enrichment experiments with ibuprofen instead of diclofenac, almost identical bacteria were isolated (as reflected from partial 16S
rRNA gene sequence; data not shown). In addition, shifts in
microbial community structure were very similar to the diclofenac-enriched community, as the same distinctive T-RF
occurred in the presence of both pharmaceuticals when other carbon sources were absent (data not shown).
We performed T-RFLP analyses of the isolated strains to
identify specific T-RFs in the total community profiles. These
only represented very minor T-RFs (with low relative peak
heights), suggesting that the isolated bacteria are not the
most abundant in polluted environments. None of the isolates corresponded to the distinctive T-RF of 292 bp, enriched
at the end of incubation with diclofenac as the only carbon
source. It seems that bacteria in the activated sludge representing this T-RF cannot be isolated by the method applied
here and alternative approaches will need to be devised in
the future. Nevertheless, we isolated seven strains capable
of growing on minimal medium with diclofenac as a sole carbon source and these represent potential models that can be
applied to better understand the metabolism of diclofenac
and study their potential for its removal from wastewater in
monoculture and in more complex communities, such as activated sludge.
Since pharmaceuticals are usually present in wastewaters
at concentrations of μg/L or less, it is possible that they do not
serve as the sole carbon sources for the wastewater microorganisms and degradation in the presence of high concentrations of many different carbon sources is probably more relevant for the degradation of diclofenac in WWTPs. Gröning
et al. (36) have proposed that the apparent lack of degradative potential of diclofenac as well as the failure to detect an
enrichment of diclofenac-depleting microbial activity both
indicate a co-metabolic nature of diclofenac transformation.
Similarly, Pieper (37) suggested that polychlorinated biphenyl (PCB)-degrading organisms do not usually use PCBs as an
energy source, but rather catabolize these substrates co-metabolically. In fact, there are a few reports in the recent years
about the degradation of diclofenac: some indicate whiterot fungi as biodegraders (38), while the first bacterium capable of using diclofenac as its sole carbon source has been
408
reported only recently (39). They showed that the isolated
strain Brevibacterium sp. D4 could biodegrade 35 % of diclofenac as a sole carbon source, while periodic feeding with
acetate as a supplementary carbon source enhanced biodegradation to the levels up to 90 %. Similarly, it was demonstrated that the isolated strain Enterobacter hormaechei D15 could
biodegrade diclofenac at an elimination rate of 52.8 %, while
in the presence of glucose as a supplementary carbon source,
degradation rate increased to approx. 82 % (40). Our preliminary assays of diclofenac degradation by the isolated bacteria have similarly indicated better elimination rates when
tests were performed in the presence of other carbon sources in addition to diclofenac. This might be a consequence of
the co-metabolic nature of diclofenac degradation or merely
the lower growth rate of the isolates when only diclofenac is
available as a carbon source. Evidently, co-metabolism of diclofenac is more efficient than its degradation as a sole carbon source and search for bacteria in a consortium able to
degrade it co-metabolically should be adapted accordingly.
CONCLUSIONS
Our results indicate that nitrification process in small fed-batch reactors was efficient and was not detectably affected by relatively high concentrations of diclofenac (up to 5
mg/L), suggesting that the selected recalcitrant and commonly detected pharmaceutical does not influence the sensitive and important process of wastewater treatment. Moreover, bacterial communities adapted to 0.2 mg/L diclofenac
did not strongly change at higher diclofenac concentrations
(5 mg/L), at least under the conditions applied in our experiment, which could explain the low diclofenac degradation
efficiency often observed in wastewater treatment plants. Influence of diclofenac at concentrations of 0.2 and 5 mg/L on
the structure of activated sludge microbial communities was
only evident when other carbon sources were not added. The
isolated strains obtained in this study are very important as
potential model bacteria to study bacterial diclofenac degradation. However, the disagreement between the bacterial community structure determined by the cultivation-independent approach and the bacterial isolation needs to be
taken into account.
ACKNOWLEDGEMENTS
The authors are grateful to Petra Zeme and Simona Leskovec for their technical assistance.
FUNDING
The authors acknowledge the financial support from
the state budget by the Slovenian Research Agency (project Z1-0776: Influence of pharmaceutical residues on activated sludge microbial communities in a pilot wastewater
treatment plant, and program P4-0116: Microbiology and biotechnology of food and environment).
October-December 2020 | Vol. 58 | No. 4
Food Technol. Biotechnol. 58 (4) 402–410 (2020)
CONFLICT OF INTEREST
The authors declare no conflicts of interest.
AUTHORS’ CONTRIBUTION
B.K. designed and performed the experiments, processed
and interpreted the data, and prepared the manuscript. B.K.
and I.M.M. revised the manuscript.
ORCID ID
B. Kraigher https://orcid.org/0000-0002-3345-272X
I. Mandic-Mulec https://orcid.org/0000-0002-7629-3255
REFERENCES
1. Daughton CG, Ternes TA. Pharmaceuticals and personal care products in the environment: Agents of subtle
change? Environ Health Perspect. 1999;107(Suppl 6):907–
38.
https://doi.org/10.1289/ehp.99107s6907
2. Kasprzyk-Hordern B, Dinsdale RM, Guwy AJ. The occurrence of pharmaceuticals, personal care products, endocrine disruptors and illicit drugs in surface water in South
Wales, UK. Water Res. 2008;42(13):3498–518.
https://doi.org/10.1016/j.watres.2008.04.026
3. Ziylan A, Ince NH. The occurrence and fate of anti-inflammatory and analgesic pharmaceuticals in sewage and fresh
water: Treatability by conventional and non-conventional
processes. J Hazard Mater. 2011;187(1–3):24–36.
https://doi.org/10.1016/j.jhazmat.2011.01.057
4. Ashton D, Hilton M, Thomas KV. Investigating the environmental transport of human pharmaceuticals to streams in
the United Kingdom. Sci Total Environ. 2004;333(1–3):167–
84.
https://doi.org/10.1016/j.scitotenv.2004.04.062
5. Kosjek T, Heath E, Kompare B. Removal of pharmaceutical
residues in a pilot wastewater treatment plant. Anal Bioanal Chem. 2007;387(4):1379–87.
https://doi.org/10.1007/s00216-006-0969-1
6. Langenhoff A, Inderfurth N, Veuskens T, Schraa G, Blokland
M, Kujawa-Roeleveld K, Rijnaarts H. Microbial removal of
the pharmaceutical compounds ibuprofen and diclofenac
from wastewater. BioMed Res Int. 2013;2013:Article ID
325806.
https://doi.org/10.1155/2013/325806
7. Quintana JB, Weiss S, Reemtsma T. Pathways and metabolites of microbial degradation of selected acidic pharmaceutical and their occurrence in municipal wastewater treated by a membrane bioreactor. Water Res.
2005;39(12):2654–64.
https://doi.org/10.1016/j.watres.2005.04.068
8. Carballa M, Omil F, Lema JM. Comparison of predicted and
measured concentrations of selected pharmaceuticals,
fragrances and hormones in Spanish sewage. Chemosphere. 2008;72(8):1118–23.
https://doi.org/10.1016/j.chemosphere.2008.04.034
9. Farré ML, Ferrer I, Ginebreda A, Figueras M, Olivella L, Tirapu L, et al. Determination of drugs in surface water and
wastewater samples by liquid chromatography–mass
spectrometry: Methods and preliminary results including toxicity studies with Vibrio fischeri. J Chromatogr A.
2001;938(1–2):187–97.
https://doi.org/10.1016/S0021-9673(01)01154-2
10. Gómez MJ, Martínez Bueno MJ, Lacorte S, Fernández-Alba
AR, Agüera A. Pilot survey monitoring pharmaceuticals and
related compounds in a sewage treatment plant located on
the Mediterranean coast. Chemosphere. 2007;66(6):993–
1002.
https://doi.org/10.1016/j.chemosphere.2006.07.051
11. Zhang Y, Geißen SU, Gal C. Carbamazepine and diclofenac:
Removal in wastewater treatment plants and occurrence
in water bodies. Chemosphere. 2008;73(8):1151–61.
https://doi.org/10.1016/j.chemosphere.2008.07.086
12. Heberer T, Reddersen K, Mechlinski A. From municipal sewage to drinking water: Fate and removal of pharmaceutical
residues in the aquatic environment in urban areas. Water
Sci Technol. 2002;46(3):81–8.
https://doi.org/10.2166/wst.2002.0060
13. Oaks JL, Gilbert M, Virani MZ, Watson RT, Meteyer CU, Rideout BA, et al. Diclofenac residues as the cause of vulture population decline in Pakistan. Nature. 2004;427(6975):630–3.
https://doi.org/10.1038/nature02317
14. Taggart MA, Cuthbert R, Das D, Sashikumar C, Pain DJ, Green RE, et al. Diclofenac disposition in Indian cow and goat
with reference to Gyps vulture population declines. Environ
Pollut. 2007;147(1):60–5.
https://doi.org/10.1016/j.envpol.2006.08.017
15. Triebskorn R, Casper H, Heyd A, Eikemper R, Köhler HR,
Schwaiger J. Toxic effects of the non-steroidal anti-inflammatory drug diclofenac: Part II. Cytological effects in liver,
kidney, gills and intestine of rainbow trout (Oncorhynchus
mykiss). Aquat Toxicol. 2004;68(2):151–66.
https://doi.org/10.1016/j.aquatox.2004.03.015
16. Paje ML, Kuhlicke U, Winkler M, Neu T. Inhibition of lotic biofilms by diclofenac. Appl Microbiol Biotechnol. 2002;59(4–
5):488–92.
https://doi.org/10.1007/s00253-002-1042-4
17. Dastidar SG, Ganguly K, Chaudhuri K, Chakrabarty AN. The
anti-bacterial action of diclofenac shown by inhibition of
DNA synthesis. Int J Antimicrob Agents. 2000;14(3):249–51.
https://doi.org/10.1016/S0924-8579(99)00159-4
18. Kraigher B, Kosjek T, Heath E, Kompare B, Mandic-Mulec I.
Influence of pharmaceutical residues on the structure of
activated sludge bacterial communities in wastewater treatment bioreactors. Water Res. 2008;42(17):4578–88.
https://doi.org/10.1016/j.watres.2008.08.006
October-December 2020 | Vol. 58 | No. 4 409
B. KRAIGHER and I. MANDIC-MULEC: Influence of Diclofenac on Sludge Bacterial Communities
19. Kraigher B, Mandic-Mulec I. Nitrification activity and community structure of nitrite-oxidizing bacteria in the bioreactors operated with addition of pharmaceuticals. J Hazard
Mater. 2011;188(1–3):78–84.
https://doi.org/10.1016/j.jhazmat.2011.01.072
20. Heuer H, Smalla K. Application of denaturing gradient gel
electrophoresis (DGGE) and temperature gradient gel electrophoresis (TGGE) for studying soil microbial communities. In: Van Elsas JD, Trevors JT, Wellington EMH, editors.
Modern soil microbiology. New York, USA: Marcel Dekker;
1997. pp. 353–73.
21. GeneScan® reference guide. Chemistry reference for the
ABI PRISM® 310 genetic analyser. Applied Biosystems, Geel,
Belgium; 2000.
Available from: https://assets.thermofisher.com/TFS-Assets/LSG/manuals/cms_040961.pdf.
22. BioNumerics v. 6.6, Applied Maths NV, Sint-Martens-Latem,
Belgium; 2011. Available from: https://download.applied-maths.com/sites/default/files/download/bn_quickguide_0.pdf.
23. Suzuki MT, Giovannoni SJ. Bias caused by template annealing in the amplification of mixtures of 16S rRNA genes by
PCR. Appl Environ Microbiol. 1996;62(2):625–30.
https://doi.org/10.1128/AEM.62.2.625-630.1996
24. Altschul SF, Gish W, Miller W, Myers EW, Lipman DJ. Basic
local alignment search tool. J Mol Biol. 1990;215(3):403–10.
https://doi.org/10.1016/S0022-2836(05)80360-2
25. Tamura K, Dudley J, Nei M, Kumar S. MEGA4: Molecular Evolutionary Genetics Analysis (MEGA) Software Version 4.0.
Mol Biol Evol. 2007;24(8):1596–9.
https://doi.org/10.1093/molbev/msm092
26. Ausec L, Kraigher B, Mandic-Mulec I. Differences in the
activity and bacterial community structure of drained
grassland and forest peat soils. Soil Biol Biochem.
2009;41(9):1874–81.
https://doi.org/10.1016/j.soilbio.2009.06.010
27. Ausec L, Zakrzewski M, Goesmann A, Schlüter A, Mandic-Mulec I. Bioinformatic analysis reveals high diversity
of bacterial genes for laccase-like enzymes. PLoS ONE.
2011;6(10):e25724.
https://doi.org/10.1371/journal.pone.0025724
28. Chen YL, Wang CH, Yang FC, Ismail W, Wang PH, Shih CJ, et
al. Identification of Comamonas testosteroni as an androgen degrader in sewage. Sci Rep. 2016;6:35386.
https://doi.org/10.1038/srep35386
29. Horinouchi M, Hayashi T, Yamamoto T, Kudo T. A new
bacterial steroid degradation gene cluster in Comamonas testosteroni TA441 which consists of aromatic-compound degradation genes for seco-steroids and 3-ketosteroid dehydrogenase genes. Appl Environ Microbiol.
2003;69(8):4421–30.
https://doi.org/10.1128/AEM.69.8.4421-4430.2003
30. Boon N, Goris J, De Vos P, Verstraete W, Top EM. Bioaugmentation of activated sludge by an indigenous
410
3-chloroaniline-degrading Comamonas testosteroni strain, I2gfp. Appl Environ Microbiol. 2000;66(7):2906–13.
https://doi.org/10.1128/AEM.66.7.2906-2913.2000
31. Qiu X, Wu P, Zhang H, Li M, Yan Z. Isolation and characterization of Arthrobacter sp. HY2 capable of degrading a
high concentration of p-nitrophenol. Bioresour Technol.
2009;100(21):5243–8.
https://doi.org/10.1016/j.biortech.2009.05.056
32. Suárez-Ojeda ME, Montón H, Roldán M, Martín-Hernández M, Pérez J, Carrera J. Characterization of a p-nitrophenol-degrading mixed culture with an improved methodology of fluorescence in situ hybridization and confocal
laser scanning microscopy. J Chem Technol Biotechnol.
2011;86(11):1405–12.
https://doi.org/10.1002/jctb.2644
33. Selvakumaran S, Kapley A, Kashyap SM, Daginawala HF,
Kalia VC, Purohit HJ. Diversity of aromatic ring-hydroxylating dioxygenase gene in Citrobacter. Bioresour Technol.
2011;102(7):4600–9.
https://doi.org/10.1016/j.biortech.2011.01.011
34. Patrauchan MA, Oriel PJ. Degradation of benzyldimethylalkylammonium chloride by Aeromonas hydrophila sp. K. J
Appl Microbiol. 2003;94(2):266–72.
https://doi.org/10.1046/j.1365-2672.2003.01829.x
35. Reardon KF, Mosteller DC, Bull Rogers JD. Biodegradation
kinetics of benzene, toluene, and phenol as single and
mixed substrates for Pseudomonas putida F1. Biotechnol
Bioeng. 2000;69(4):385–400.
https://doi.org/10.1002/1097-0290(20000820)69:4<385::A
ID-BIT5>3.0.CO;2Q
36. Gröning J, Held C, Garten C, Claußnitzer U, Kaschabek SR,
Schlömann M. Transformation of diclofenac by the indigenous microflora of river sediments and identification of a
major intermediate. Chemosphere. 2007;69(4):509–16.
https://doi.org/10.1016/j.chemosphere.2007.03.037
37. Pieper DH. Aerobic degradation of polychlorinated biphenyls. Appl Microbiol Biotechnol. 2005;67(2):170–91.
https://doi.org/10.1007/s00253-004-1810-4
38. Rodarte-Morales AI, Feijoo G, Moreira MT, Lema JM.
Biotransformation of three pharmaceutical active compounds by the fungus Phanerochaete chrysosporium in a
fed batch stirred reactor under air and oxygen supply. Biodegradation. 2012;23(1):145–56.
https://doi.org/10.1007/s10532-011-9494-9
39. Bessa VS, Moreira IS, Tiritan ME, Castro PML. Enrichment of
bacterial strains for the biodegradation of diclofenac and
carbamazepine from activated sludge. Int Biodeter Biodegr. 2017;120:135–42.
https://doi.org/10.1016/j.ibiod.2017.02.008
40. Aissaoui S, Ouled-Haddar H, Sifour M, Harrouche K, Sghaier
H. Metabolic and co-metabolic transformation of diclofenac by Enterobacter hormaechei D15 isolated from activated sludge. Curr Microbiol. 2017;74(3):381–8.
https://doi.org/10.1007/s00284-016-1190-x
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