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Removal of pharmaceuticals in aerated biofilterswith manganese feeding
Yongjun Zhang a,*, Hong Zhu b, Ulrich Szewzyk b, Sven Uwe Geissen a
a Technische Universit€at Berlin, Chair of Environmental Process Engineering, Sekr. KF2, Strasse des 17.Juni 135,
10623 Berlin, Germanyb Technische Universit€at Berlin, Chair of Environmental Microbiology, Sekr. BH 6-1, Ernst-Reuter-Platz 1, 10587
Berlin, Germany
a r t i c l e i n f o
Article history:
Received 31 May 2014
Received in revised form
16 December 2014
Accepted 7 January 2015
Available online 14 January 2015
Keywords:
Biodegradation
Biofiltration
Tertiary treatment
Emerging pollutant
Biofilm
* Corresponding author. Tel.: þ49 (0)30 31425E-mail address: yongjun.zhang@tu-berlin
http://dx.doi.org/10.1016/j.watres.2015.01.0090043-1354/© 2015 Elsevier Ltd. All rights rese
a b s t r a c t
A tertiary treatment step is required in current wastewater treatment plants to remove
trace pollutants and thus to prevent their extensive occurrence in the aquatic environ-
ment. In this study, natural MnOx ore and natural zeolite were separately used to pack two
lab-scale aerated biofilters, which were operated in approximately 1.5 years for the
removal of frequently occurring pharmaceuticals, including carbamazepine (CBZ), diclo-
fenac (DFC), and sulfamethoxazole (SMX), out of synthetic and real secondary effluents.
Mn2þ was added in the feeds to promote the growth of iron/manganese oxidizing bacteria
which were recently found to be capable of degrading recalcitrant pollutants. An effective
removal (80e90%) of DFC and SMX was observed in both biofilters after adaptation while a
significant removal of CBZ was not found. Both biofilters also achieved an effective removal
of spiked Mn2þ, but a limited removal of carbon and nitrogen contents. Additionally, MnOx
biofilter removed 50% of UV254 from real secondary effluent, indicating a high potential on
the removal of aromatic compounds.
© 2015 Elsevier Ltd. All rights reserved.
1. Introduction
Various pharmaceutical active compounds (PhACs) have been
detected in water bodies. Although at trace concentrations,
they still have a potential impact on the aquatic ecological
system (Brausch et al., 2012). This impact could become
complex within the interactions of various species. For
instance, the antibiotics would lead to an increased resistance
ofmicroorganisms who can spread their resistance genes into
other microbes (Baquero et al., 2008). It was also reported that
the dramatic population decline of vultures in the India sub-
continent was resulted from the residue of diclofenac (DFC) in
298; fax: þ49 (0)30 31425.de (Y. Zhang).
rved.
livestock bodies which held an important position in the food
chain of vultures (Green et al., 2004). Main sources of PhACs
are the human and veterinary applications. PhACs from the
former source would mostly enter wastewater treatment
plants (WWTPs) via a sewage system while those from the
latter are normally diffused. Therefore, one important strat-
egy to control the contamination of PhACs is to upgrade the
current WWTPs which currently cannot effectively eliminate
all PhACs (Zhang and Geißen, 2010).
Many technologies have been studied for the removal of
PhACs, including AOP, ozonation, membrane separation,
sorption, etc (Basile et al., 2011). Among them, ozonation and
activated carbon adsorption are most promising and have
487.
wat e r r e s e a r c h 7 2 ( 2 0 1 5 ) 2 1 8e2 2 6 219
been installed at full scale (Reungoat et al., 2010;
Zimmermann et al., 2011). However, those chemical/physical
methods mostly require high investment and/or operation
costs. Therefore, the development of a cost-efficient method
is of great concern.
Iron/manganese oxidizing bacteria (IMOB) can mediate the
oxidation of Fe2þ and Mn2þ to their high-valency states and
thus play an important role in the natural cycle of iron and
manganese (Emerson et al., 2010). IMOBwidely exist in natural
water bodies and technical water systems. For example,
Cerrato et al. (2010) detected several strains of IMOB in the
sediment of surface water, on the filter materials of drinking
water treatment, and in thewater distribution pipes evenwith
the presence of chlorine. IMOB have been widely involved in
the removal of Fe2þ and Mn2þin the drinking water treatment
(Gouzinis et al., 1998).
Recently, researchers found that IMOB are also capable of
degrading PhACs. One degradation mechanism is the in vitro
adsorption and oxidation by the biogenic manganese oxides
(Bio-MnOx). Sabirova et al. (2008) studied the removal of 17a-
ethinylestradiol (EE2) with several IMOB strains and they
found that IMOB can effectively degrade EE2with the presence
of Mn2þ. When a bacteriostatic agent (sodium azide) was
applied to inhibit the manganese-oxidizing activity of IMOB,
the authors still observed 80% removal of EE2. Thus they
concluded that the degradation was initiated by the attack of
biogenic manganese oxides (Bio-MnOx). This is in accordance
with the fact that manganese oxides hold a high redox po-
tential and both synthetic and biogenic MnOx are able to
oxidize many organic and inorganic pollutants (Hennebel
et al., 2009; Zaman et al., 2009; Zhang et al., 2008a). However,
it was also found that the participation of active IMOB was
beneficial and sometimes essential for pollutant oxidation.
Murray and Tebo (2007) found that IMOB Bacillus sp. strain SG-
1 can accelerate the oxidation of Cr3þ compared to both syn-
thetic and biogenicMnOx. Forrez et al. (2010) reported that the
addition of either sodium azide or lysozyme (a cell lysis agent)
resulted in a significant inhibition of DFC oxidation by Bio-
MnOx of IMOB Pseudomonas putida. Meerburg et al. (2012)
further reported that the heat inactivated Bio-MnOx of P.
putida cannot remove DFC. Our previous study also observed
that several IMOB strains can remove DFC but the heat inac-
tivated biomass resulted in no removal (Zhu et al., 2012).
Therefore, other mechanisms may exit to initiate the
pollutant removal which occurs with the active manganese-
oxidizing metabolism of IMOB, e.g. via the involvement of a
ligand-bound Mn3þ intermediate (Meerburg et al., 2012).
Therefore, for real applications, the active involvement of
IMOB should be guaranteed.
However, only few of studies have been published on the
removal PhACs with IMOB in technical systems. De Rudder
et al. (2004) used a MnO2 mineral as a packing material in a
filter for the removal of EE2 from tap water. They noticed that
the loaded EE2 on MnO2 was significantly beyond its adsorp-
tion capacity and thus they proposed that a self-regeneration
process may occur. Forrez et al. (2009) compared two filters
packed with the same mineral and plastic polyethylene rings.
The filter with plastic rings was additionally fed with Mn2þ
and therefore the rings were expected to be covered with Bio-
MnOx. Both filters can successfully remove EE2 but the
biofilter with plastic rings failed to recover the effective
removal after an increase of EE2 loading and also resulted in a
higher Mn2þ level in effluent than the filter with MnO2 min-
eral. Recently, Forrez et al. (2011) added Bio-MnOx
(160 mg Mn4þ L�1) into the membrane module of a biore-
actor and achieved an effective removal of many PhACs.
However, such a configuration is complex to operate and may
have some negative impacts on the membrane performance,
e.g. fouling, permeability, etc.
In the present study, aerated biofilters were studied, with
natural MnOx ore and zeolite as packing materials. Mn2þ was
additionally spiked into the feeding wastewater to promote
the growth of IMOB. Therefore, a simple configuration and
active IMOB were expected with the studied aerated biofilters
and the removal of three widely detected PhACs, including
DFC, carbamazepine (CBZ) and sulfamethoxazole (SMX), was
investigated with both synthetic and real wastewater.
2. Material and methods
2.1. Materials
MnOx ore (Aqua-mandix®) and natural zeolite (Bigadia®) were
supplied by Aqua-Techniek, Netherlands, and EgeZeolit,
Turkey, respectively. Their chemical compositions and phys-
ical properties are shown in Table 1. Diclofenac sodium, car-
bamazepine, sulfamethoxazole (all with analytical grade), and
manganese(II) sulfate monohydrate with ReagentPlus® grade
were purchased from Sigma Aldrich, Germany.
2.2. Biofilter and operation
The MnOx ore and zeolite were used to pack two biofilters,
named Mn-Biofilter and Zeo-Biofilter, respectively. Both bio-
filters were constructed out of cylindrical acrylic glass
(diameter 7 cm; effective height 45 cm; packed height 35 cm)
andwere aeratedwith a fine-bubblemembrane diffuser at the
bottom (Fig. 1). Aluminum foil was used to wrap the filters to
prevent light. The aeration rate was regulated with a needle
valve at approximately 200 mL h�1 and 3e5 mg L�1 dissolved
oxygen was achieved in both biofilters. The biofilters were fed
with wastewater (compositions described below) via a multi-
channel peristaltic pump at 430 mL d�1, which resulted in
an empty bed contacting time (EBCT) of 28 h and 20 h in Mn-
Biofilter and Zeo-Biofilters, respectively. In order to avoid the
clogging problem, another multi-channel peristaltic pump
was used to recycle water in both biofilters at 300 mL h�1 to
generate a superficial velocity of 3 and 2 m h�1 in Mn-Biofilter
and Zeo-Biofilters, respectively.
Both synthetic wastewater and effluent from a localWWTP
were applied to feed the biofilters. The synthetic wastewater
was produced via diluting the OECD composition (OECD, 2001)
to simulate a WWTP effluent: 1-L tap water containing
peptone 16mg,meat extract 11mg, NH4Cl 5mg, NaNO3 10mg,
K2HPO4 2.8 mg, NaCl 0.7 mg, CaCl2$2H2O 0.4 mg, Mg2SO4$7H2O
0.2 mg (DOC 10 mg L�1, TN 9 mg L�1, pH 7.5). The collected
WWTP effluent containedDOC 11 ± 3mg L�1, TN 15 ± 2mg L�1,
TP 0.6 mg L�1, conductivity 1260 ± 50 mS cm�1, UV254 0.3 cm�1,
pH 7.0. In addition, 20 mg L�1 Mn2þ was added into both
Table 1 e Chemical composition and physical properties of MnO2 and zeolite used in the biofilters.
MnO2 Composition MnO2 78%, Fe2O3 6.2%, SiO2 5.2%, Al2O3 3.1%
Size 0.5e1.0 mm
Bulk density 2000 kg m�3
Void ratio 0.3
Zeolite Composition SiO2 70.14%, Al2O3 11.46%, K2O 4.37%, CaO 2.98%, MgO 1.46%, Fe2O3 0.85%
Size 2e5 mm
Bulk density 1000 kg m�3
Void ratio 0.4
wat e r r e s e a r c h 7 2 ( 2 0 1 5 ) 2 1 8e2 2 6220
feedings. SMX, CBZ and DFCwere also spiked in both feedings
at concentrations described below.
The biofilters were initially filled with a solution of 10-d old
IMOB strains (Leptothrix sp. and Pseudomonas sp.) which were
separately cultivated in a liquid medium (lepthothrix strains
media 2) prepared according to (Atlas, 1997). After 2 d, the
biofilters were continuously fed with the synthetic
wastewater.
Both biofilters experienced the following operation phases:
A) starting with 2 mg L�1 PhACs in synthetic wastewater; B)
working phase with the same feeding; C) feeding with
1.5 mg L�1 PhACs in synthetic wastewater; D) feeding with
1.5 mg L�1 extra PhACs in WWTP effluent; E) feeding with
1.5 mg L�1 extra PhACs in WWTP effluent without feeding
Mn2þ. The final concentrations of SMX, CBZ andDFC inWWTP
effluent were 2, 5, 4 mg L�1, respectively. The biofilters were
assumed to enter the working phase (Phase B) when an
effective SMX removal was observed. A transition period of
2e4 weeks occurred between the other sequent operation
phases.
3 samples were taken per week in the operation Phase A
and B and 2 samples in the other phases. After filtration with
0.45 mm cellulose nitrate membrane, all samples were stored
at 4 �C for the collective analysis.
2.3. Control tests
Some carrier materials were taken from both biofilters and a
defined amount (wet weight: 10 g Mn ore, 20 g zeolite) was
Fig. 1 e The scheme of an aerated biofilter.
added into 50 mLWWTP effluent spiked with 2 mg L�1 PhACs.
A control test with the presence of sodium azide (50 mg L�1)
was also conducted for bothmaterials. Samples of water were
taken in intervals during 10 d and analyzed with HPLC after
centrifugation.
2.4. Sample analysis
PhACs at the high concentrationwere analyzedwith HPLC-UV
(Agilent 1200) system equippedwith a Gemini-NX C18 column
(250 � 4.6 mm, 5 mm) and a SecurityGuard pre-column. The
column temperature was set at 30 �C. The injection volume
was 50 mL. Elution was conducted with Millipore water (0.1%
formic acid) and acetonitrile, at a flow rate of 1.0 mL min�1
with the following gradient: 40% acetonitrile at 0 min, 60% at
8 min, 90% at 9 min, 60% at 14 min, and 40% at 15 min (the
end). The eluted SMX, CBZ and DFC were detected by a UV
detector at 265 nm, 285 nm, and 275 nm, respectively. The
limit of quantification was 50 mg L�1 for all three PhACs.
The samples with low concentrations of PhACs were
analyzed with a HPLC-MS system, consisting of Agilent 1260
HPLC and AB SciexQTrap 5500 MS with an electrospray ioni-
zation interface. Analysis was carried out in positive ion
mode. A Synergi Hydro-RP column (150 � 3 mm) with a Se-
curity Guard column AQ C18 was used for the chromato-
graphic separation. 70 mL of sample was injected into the
HPLC-MS system. The column oven temperature was set to
40 �C. The elution solution consisted of Millipore water with
0.1% formic acid and 10 m Mammonium format (A) and
methanol with 0.1% formic acid (B). The gradient elution
program was set as follows: 0e4 min, 98% (A); 5 min, 60%;
16.5 min, 0%; 22 min, 0%; 22.1 min, 98%. The run time was
28 min and the flow rate was 0.50 mL min�1.
Dissolved organic carbon (DOC) and total Nitrogen (TN)
were analyzedwith aDOC/TN analyzer (Analytik Jenamulti N/
C 3100). DOC was measured as non-purgeable organic carbon.
Mn2þ was analyzed by an atomic absorption spectroscopy
(PerkinElmer PinAAcle 900 Z). The absorbance of ultraviolet at
254 nm (UV254) was measured with a photo spectrometer
(DR5000, Hach Lange GmbH). Total phosphorus (TP) was
measured with a cuvette test from Hach Lange GmbH.
3. Results
3.1. Removal of PhACs
Fig. 2a shows the removal efficiencies of SMX, CBZ, and DFC in
Mn-Biofilter at different operation phases. About 95% of DFC
Fig. 3 e UVevis spectra of SMX and DFC in Mn-Biofilter (a)
and Zeo-Biofilter (b).
Fig. 2 e Removal efficiencies of SMX, CBZ and DFC in Mn-
Biofilter (a) and Zeo-Biofilter (b) in different operation
phases: A) starting with 2 mg L¡1 PhACs in synthetic
wastewater; B) working phase with the same feeding; C)
feeding with 1.5 mg L¡1 PhACs in synthetic wastewater; D)
feeding with 1.5 mg L¡1 extra PhACs in WWTP effluent; E)
feeding with 1.5 mg L¡1 extra PhACs in WWTP effluent
without feeding Mn2þ. The whiskers show 5% and 95%
percentiles and the solid dots stand for the mean value.
wat e r r e s e a r c h 7 2 ( 2 0 1 5 ) 2 1 8e2 2 6 221
was removed immediately after starting the biofilter. On the
contrary, only 10% of SMX was removed in the starting period
of 3 months. However, thereafter the removal of SMX rapidly
increased to >90% in several days (defined as the start of
operation Phase B) and kept stable. The removal of both SMX
and DFC was dropped to 70% and 80% (median value),
respectively, after switching to the trace concentration (Phase
C). More interestingly, both removal efficiencies were elevated
with WWTP effluent (Phase D): 90% of SMX and 96% of DFC.
After excluding Mn2þ in the feeding (Phase E), the removal of
both SMX and DFC was slightly decreased and a broader
variation was also observed here than in the previous phase.
CBZ obtained a negative removal efficiency in Mn-Biofilter
after the trace feeding concentrations were applied, which
indicates that the removal of CBZ in Phase A and B, although
at very low efficiencies, was probably due to the sorption onto
the carrier material.
The removal of SMX, CBZ, and DFC in Zeo-Biofilter is
shown in Fig. 2b. It can be found that all three substanceswere
poorly removed in the approximately 4-month starting phase.
However, an effective removal of SMX appeared quickly after
the starting phase, just like the Mn-Biofilter. An escalation of
DFC removal was clearly observed through the operation
phases and finally a median removal of 88% was achieved in
phase D after about 1-year operation. A slightly decreased
removal of SMX (79%) and DFC (74%) was also found in Phase
E, similar to Mn-Biofilter. This biofilter also resulted in a
limited removal of CBZ in Phases A-D but it was very inter-
esting that CBZ was slightly removed (<20%) in Phase E.
UVevis spectroscopy was applied to evaluate the evolution
of aromatic contents in the treated wastewater. At the end of
Phase E, 10 mg L�1 DFC or SMX was separately fed into both
biofilters and the effluents were scanned with UVevis spec-
troscopy. The UVevis spectra of SMX and DFC are shown in
Fig. 3. It can be found that Mn-Biofilter effectively destroyed
the aromatic structure of DFC while SMX was only partially
removed (Fig. 3a). An opposite result was observed in Zeo-
Biofilter where the aromatic structure of SMX was
completely eliminated (Fig. 3b). Nevertheless, the partial
removal of DFC or SMX may be due to the high loading, it
might be possible that the aromatic structure of both
Fig. 4 e Concentration profiles of DFC and SMX
(C0 ¼ 2mg L¡1) in wastewater andwastewater plus sodium
azide with the inoculated carrier materials from Mn-
Biofilter (a) and Zeo-Biofilter (b). Error bars stand for the
standard deviation of triple tests.
Fig. 5 e Effluent (C) to influent (C0) ratios of DOC and TN in
Mn-Biofilter (a) and Zeo-Biofilter (b) with synthetic
wastewater and WWTP effluent as feedings. The whiskers
show 5% and 95% percentiles and the solid dots stand for
the mean value.
wat e r r e s e a r c h 7 2 ( 2 0 1 5 ) 2 1 8e2 2 6222
substances can also be destroyed at a trace level. However, the
transformation products were not identified in this study,
which makes the degradation pathways still unclear.
A control test was conducted in order to clarify the bio-
logical and chemical effects of the biomass on both carrier
materials. As shown in Fig. 4, the presence of sodium azide
significantly depressed the elimination of SMX by both carrier
materials, indicating that the elimination of SMX can be
attributed to the biological oxidation of manganese in both
biofilters. Azide also negatively influenced DFC removal with
the inoculated zeolite but its influence on DFC removal with
inoculated MnOx ore was limited, which indicates a signifi-
cant removal by chemical reactions.
3.2. Removal of DOC, TN, TP, and UV254
DOC and TN were monitored during the whole operation of
biofilters while TP and UV254 only when WWTP effluent was
applied. Fig. 5 shows the C/C0 ratios of DOC and TN in three
phases: feeding with synthetic wastewater plus Mn, with
WWTP effluent plusMn, andwithWWTP effluentwithoutMn.
The absolute concentrations of DOC and TN can also been
found in Supplementary Information (Fig. 1s). In general, both
biofilters resulted in a limited removal of DOC and TN. DOC
removal was slightly enhanced in both biofilters when they
were fed with WWTP effluent plus Mn: Median values of C/C0
were deceased from 83% to 64% and 92%e76% in Mn-Biofilter
and Zeo-Biofilter, respectively. However, when Mn was
excluded in the feed, the valuewas slightly increased to 68% in
Mn-Biofilter and 95% in Zeo-Biofilter. On the other hand, the
TN removal in both biofilters was dropped down after WWTP
effluent plus Mn was applied. The median C/C0 ratios of TN
were increased from 86% to 93% in Mn-Biofilter and from 73%
to 93% in Zeo-Biofilter. The median C/C0 value of TN was
decreased to ~86% in both biofilters when Mn was excluded.
The removal efficiencies of UV254 and TP are present in
Fig. 6. Zeo-Biofilter showed a moderate removal of TP with a
median value of 41% which was much higher than the
removal in Mn-Biofilter (23%). However, UV254 was more
effectively removed in Mn- than Zeo-Biofilter: 51% vs. 22% of
median value. UV254 in the original WWTP effluent was
approximately 0.3 cm�1 and median values of UV254 were 0.15
Fig. 6 e Removal of UV254 and total phosphorus (TP) in Mn-
Biofilter and Zeo-Biofilter fed with WWTP effluent. The
whiskers show 5% and 95% percentiles and the solid dots
stand for the mean value.
Fig. 7 e Mn concentrations in the effluents of Mn-Biofilter
(a) and Zeo-Biofilter (b). The whiskers in boxplot show 5%
and 95% percentiles and the solid dots stand for the mean
value.
wat e r r e s e a r c h 7 2 ( 2 0 1 5 ) 2 1 8e2 2 6 223
and 0.24 cm�1 in the effluents ofMn-Biofilter and Zeo-Biofilter,
respectively. A significant influence of Mn in the feed was not
found on the removal of UV254 and TP in both biofilters.
3.3. Residual manganese
Manganese is normally not controlled for the effluent
discharge. WHO suggests a provisional value of 0.5 mg L�1
manganese for the hazardous control of drinking water while
a ten-fold lower value (50 mg L�1) is recommended by USA and
Europe to avoid odor and staining effects of manganese.
The spiked manganese was effectively removed with both
biofilters and its concentration in effluentwas generally below
the ahead mentioned standards in all operation stages as
shown in Fig. 7. However, a slightly elevatedMn concentration
was observed in the effluents of both biofilters when the real
WWTP effluent was applied to replace the synthetic waste-
water.Without spikingMn in the real wastewater, only a trace
amount of Mn was detected in both biofilters, which was in
the range of Mn content of applied WWTP effluent. It also
indicated that Mn was not released from the previously fixed
Mn on the packing materials.
4. Discussion
Current WWTPs cannot effectively remove SMX and DFC
(Larcher and Yargeau, 2011; Zhang et al., 2008b). Our results
provide an alternative solution to remove both substances
with a simple reactor and an ease operation. However, adap-
tation is required for both biofilters to effectively remove SMX:
1 month in Mn-Biofilter, 1.5 month in Zeo-Biofilter. It might be
related to the development of some SMX-degrading IMOB
since the removal of SMX should be a result of biological
manganese oxidation as shown in the control test. It is highly
possible that the Mn3þ intermediate formed during the bio-
logical oxidation of Mn2þ was involved in the degradation of
SMX, like the degradation of DFC by P. putida (Meerburg et al.,
2012). Nevertheless, a high level of SMX (2 mg L�1) was used
during the adaptation in this study. Baumgarten et al. (2011)
found that the high concentration of SMX can largely
shorten the adaptation time for SMX removal in aerobic sand
filters (HRT ~14 d). Therefore, it might be possible that both
biofilters in this study would need a longer adaption time
when starting immediately with secondary effluent where the
concentration of SMX is in the range of low mg L�1.
It took about 1 year to obtain an effective removal of DFC in
Zeo-Biofilter. It is interesting to notice that the removal of DFC
in Zeo-Biofilter was gradually increased along the operation
time, during which the spiked manganese should be biologi-
cally fixed in the form of MnOx on the zeolite (estimated as
~1.5 mg Mn per g zeolite according to the mass balance). After
the whole operation period, one can clearly found the black
biofilm on zeolite (Fig. 2s in Supplementary Information). A lot
of manganese was also detected on zeolite in the elementary
analysis with energy dispersive X-ray spectroscopy (Fig. 3s in
Supplementary Information). This Bio-MnOx on zeolite can
still eliminate DFC even with the presence of azide, however,
with a lower kinetic rate thanwithout azide as shown in Fig. 4.
It indicates that the involvement of biological Mn oxidation is
beneficial for the activity of Bio-MnOx, which is in agreement
with the study of Forrez et al. (2010).
On the other hand, Mn-Biofilter can achieve an immediate
removal of DFC, which should be attributed to the catalytic
effects of mineral MnOx, as demonstrated by the recent study
of Huguet et al. (2013). They also found that the oxidation
capacity of MnOx was gradually diminished during the reac-
tion with DFC (~20 mg L�1 in clean water). After the treatment
of 350 empty-bed volumes, the oxidation capacity dis-
appeared, which the authors attributed to the reductive
dissolution and saturation of the active sites. In this study, the
biofilters were fed with approximately 400 empty-bed
wat e r r e s e a r c h 7 2 ( 2 0 1 5 ) 2 1 8e2 2 6224
volumes of secondary effluent and a significant reduction of
DFC removal was not observed. The bioactivity in biofilters
may regenerate the active sites of MnOx or Bio-MnOx via e.g.
oxidizing the dissolute Mn2þ back to MnOx, which should be
confirmed with a separate study. Furthermore, it is worth
pointing out that the activity of Bio-MnOx is related to its
poorly crystalline layer structure (e.g. hexagonal sheet sym-
metry) (Villalobos et al., 2003). The structure of Bio-MnOxmay
change during the reaction and consequently its activity could
decrease. Recently, Learman et al. (2011) analyzed the evolu-
tion of structure and activity of Bio-MnOx from Roseobacter sp.
AzwK-3b during the in vitro oxidation ofMn2þ. They found that
78% of Bio-MnOx hexagonal symmetry was gradually trans-
ferred to a pseudo-orthogonal symmetry in 5 d and the latter
almost lost its activity of Mn2þ oxidation. It indicates that a
continuous generation of Bio-MnOx may be required to sus-
tain the catalytic activity. However, the structure evolution of
MnOx may also occur during the reaction with organic pol-
lutants, which is still not reported according to our
knowledge.
UV254 indicates the color and aromatic contents (e.g. humic
substances) in wastewater, which may be problematic, e.g. by
forming disinfection byproducts in chlorination and causing
membrane fouling (Tripathi et al., 2011; Zheng et al., 2010).
Wert et al. (2009) also found a strong correlation between
UV254 removal and pharmaceutical elimination from tertiary
treated wastewater by O3 and O3/H2O2. In their study,
4e8 mg L�1 of ozone dosage was required to remove 50% of
UV254. However, UV254 cannot be effectively removed by most
biofilters. For example, the slow sand filters can only remove
5e6% of UV254 from a secondary effluent (Zheng et al., 2010).
Wang et al. (2008) combined an ozonation column and a bio-
filter with a clay media to treat a WWTP secondary effluent.
They obtained 60% and 15% removal of UV254 in ozonation
(8 mg L�1 dosage) and biofilter, respectively. Lee et al. (2012)
applied a MBR to treat a WWTP secondary effluent, which
was followed by an ozonation column and a biofilter packed
with anthracite coal at EBCT of 20min. They found that 56% of
UV254 in theMBR effluent was removedwith ozone at a dosage
of 8 mg L�1 (Lee et al., 2012). The subsequent biofilter removed
negligible UV254. In the present study, about 50% of UV254 was
removed in Mn-Biofilter without the complex operation. It
indicates that Mn-Biofilter might also be applied as a pre-
treatment for membrane separation and drinking water
chlorination.
The low removal of carbon contents in both biofilters is
in accordance with other biofilters for the treatment of
secondary effluent: e.g. 25% COD removal (Wang et al.,
2008), 20% of DOC removal of pre-ozonated effluent
(8 mg L�1 ozone dosage) (Lee et al., 2012). Both biofilters had
a sufficient oxygen supply (dissolved oxygen 3e5 mg L�1)
and thus the denitrification cannot significantly occur.
Therefore, the limited removal of TN was found in both
biofilters. Nevertheless, a higher removal of DOC can be
noticed in Mn-Biofilter than Zeo-Biofilter (Fig. 5). It may be
due to the oxidation of humics by MnOx. It has been found
that chemically synthesized MnOx can mediate the oxida-
tion of humic substances to generate low-molecular-weight
organic compounds, including pyruvate, acetone, formal-
dehyde, etc. (Sunda and Kieber, 1994). That also partially
explained the better removal of UV254 by Mn-Biofilter.
However, the presence of humic substances could inhibit
the oxidation of organic pollutants. Li et al. (2012) found that
humic acid can dramatically decrease the mineralization of
phenolic monomers (catechol and p-coumaric acid) by
chemical MnO2.
It is also worth mentioning that the material cost of MnOx
ore per volumetric unit is more expensive (3e5 times) than
granular activated carbon. Therefore, the initial investment
for a MnOx filter should also be higher. However, activated
carbon is usually suffered from the declined adsorption ca-
pacity and consequently decreased removal of organics. Mn-
Biofilter did not encounter this problem. It is also possible
that activated carbon can further remove recalcitrant organics
thanks to the biofilm (Reungoat et al., 2012). However, the
development of mature biofilm may takes a long term, up to
several years. The natural zeolite is a less expensive material
but it took Zeo-Biofilter a long adaption term to degradeDFC in
this study. Its adaption may be shorten, e.g. by increasing the
loading ofmanganese, which surely should be confirmedwith
another study.
5. Conclusions and outlook
The following conclusions can be made from this study:
1) SMX and DFC can be effectively removed in the aerated
biofilter packed with natural MnOx and zeolite. However,
Zeo-Biofilter required a long adaption time (~1 year) to
remove DFC. Mn-Biofilter and Zeo-Biofilter can also effec-
tively eliminate the aromatic contents of DFC and SMX,
respectively.
2) DFC was mainly removed by the oxidation of MnOx while
SMX during the biological oxidation of manganese. The
bioactivity was also beneficial for the removal of DFC by
MnOx.
3) Spiked Mn2þ can be effectively removed and therefore
environmental impact of manganese feeding is minimal.
4) Both biofilters resulted in a limited removal of DOC and TN.
However, Mn-Biofilter obtained ~50% removal of UV254
from the secondary effluent, indicating a high potential of
eliminating aromatic contents in wastewater.
Further studies are required to shorten the adaption time
of Zeo-Biofilter, e.g. via increasing the loading of manganese
and to identify the crystal structure of Bio-MnOx and its in-
fluence on the elimination of DFC.
Acknowledgment
This study is part of the project TransRisk (grant No.
02WRS1276J) financed by the German BMBF within the
program of RiskWa. We also appreciate the work of Dr.
Wolfram Seitz, Zweckverband Landeswasserversorgung at
Langenau and Ms. Elena Kaiser, Bundesanstalt fur Gew€as-
serkunde (BfG) at Koblenz, for the analysis of
pharmaceuticals.
wat e r r e s e a r c h 7 2 ( 2 0 1 5 ) 2 1 8e2 2 6 225
Appendix A. Supplementary data
Supplementary data related to this article can be found at
http://dx.doi.org/10.1016/j.watres.2015.01.009.
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