ORIGINAL PAPER

The impact of titanium dioxide nanoparticles on biologicalnitrogen removal from wastewater and bacterial communityshifts in activated sludge

Dapeng Li • Fuyi Cui • Zhiwei Zhao •

Dongmei Liu • Yongpeng Xu • Huiting Li •

Xiaonan Yang

Received: 18 December 2012 / Accepted: 3 May 2013

� Springer Science+Business Media Dordrecht 2013

Abstract The potential impact of titanium dioxide

nanoparticles (TiO2 NPs) on nitrogen removal from

wastewater in activated sludge was investigated using a

sequencing batch reactor. The addition of 2–50 mg L-1

of TiO2 NPs did not adversely affect nitrogen removal.

However, when the activated sludge was exposed to

100–200 mg L-1 of TiO2 NPs, the effluent total

nitrogen removal efficiencies were 36.5 % and

20.3 %, respectively, which are markedly lower than

the values observed in the control test (80 %). Further

studies showed that the decrease in biological nitrogen

removal induced by higher concentrations of TiO2 NPs

was due to an inhibitory effect on the de-nitrification

process. Denaturing gradient gel electrophoresis pro-

files showed that 200 mg L-1 of TiO2 NPs significantly

reduced microbial diversity in the activated sludge. The

effect of light on the antibacterial activity of TiO2 NPs

was also investigated, and the results showed that the

levels of TiO2-dependent inhibition of biological nitro-

gen removal were similar under both dark and light

conditions. Additional studies revealed that different

TiO2 concentrations had a significant effect on dehy-

drogenase activity, and this effect was most likely the

result of decreased microbial activity.

Keywords Titanium dioxide nanoparticles �Biological nitrogen removal � Bacterial community

shift � Activated sludge

Introduction

Nanoparticles (NPs) are defined as particles with novel

and distinctive physicochemical properties the size of

which in the range of 1–100 nm in at least one

dimension (Julia et al. 2009; Moore 2006). The

importance of nanotechnology in modern society is

demonstrated by the broad range of scientific and

technological applications that use NPs (Julia et al.

2009; Moore 2006; Nel et al. 2006), including imaging

and medical devices, cosmetics, fabrics, health prod-

ucts, and water remediation technologies (Warren

et al. 1998). Although nanotechnology holds great

potential for valuable environmental uses, the rapid

increase in the number of nanotechnology-enhanced

products raises concerns regarding the potential

adverse effects of these particles on human health

and the environment (Choi and Hu 2008; Giammar

et al. 2007; Kim et al. 2007; Zhang and Karn 2005;

Maynard et al. 2006). Many studies have investigated

the effect of NPs on the environment and attempted to

predict their environmental concentrations (Zheng

et al. 2011; Nowack and Bucheli 2007; Lv et al. 2008),

and the increasing use of nanoparticle-containing

products has resulted in the release of NPs into

wastewater treatment plants (WWTPs) (Blaster et al.

D. Li � F. Cui (&) � Z. Zhao � D. Liu � Y. Xu �H. Li � X. Yang

State Key Laboratory of Urban Water Resource and

Environment, Harbin Institute of Technology,

73 Huanghe Road, Nangang District,

Harbin 150090, China

e-mail: hit_cuifuyi@hotmail.com

123

Biodegradation

DOI 10.1007/s10532-013-9648-z

2008; Gottschalk et al. 2009). Limbach et al. (2008)

found that large amounts of cerium oxide NPs that

were released into a model WWTP were adsorbed by

activated sludge, and this was the major mechanism

for NP removal in a conventional activated sludge

system (Zheng et al. 2011). Nevertheless, few studies

have been conducted to determine whether the

absorbed NPs induce adverse effects in activated

sludge.

TiO2 is an effective opacifier and is used as a

pigment in paints, paper, inks, and plastics (Adams

et al. 2006). Due to the widespread use of TiO2,

research has been conducted on its potential toxicity

(Adams et al. 2006; Rincon and Pulgarin 2005;

Lonnen et al. 2005). In previous studies, TiO2 particles

that were toxic to bacteria ranged in size from tens of

nanometers to hundreds of micrometers. Recent

studies have confirmed that TiO2 NPs are toxic to

both Gram-negative and Gram-positive bacteria

(Block et al. 1997; Kwark and Kim 2001). Some

studies have found Gram-positive bacteria to be less

sensitive to TiO2 NPs than Gram-negative bacteria,

but other studies have observed the opposite results.

For example, (Rincon and Pulgarin 2005) reported that

Gram-positive Bacillus subtilis was less sensitive to

the effects of TiO2 than a pure culture of Gram-

negative Escherichia coli in a mixed culture experi-

ment. However, Adams et al. (2006) found that the

Gram-positive Bacillus subtilis was more sensitive to

the addition of three types of NPs (TiO2 NPs, SiO2

NPs, and ZnO NPs) than were Gram-negative E. coli.

The toxicity of TiO2 may be related to the mechanisms

by which the particles act on cells, including the

disruption of membranes or the membrane potential,

protein oxidation, genotoxicity, the interruption of

energy transduction, the formation of reactive oxygen

species, and the release of toxic constituents (Klaine

et al. 2008). TiO2 is photosensitive and produces

reactive oxygen species (ROS) in the presence of light

(Adams et al. 2006; Wei et al. 1994). During these

reactions, light of a specific wavelength is usually

provided by sunlight or high-intensity lamps. How-

ever, in one study, inhibitory effects were observed

under dark conditions, suggesting that undetermined

mechanisms may contribute to the observed toxicity.

Nevertheless, the current knowledge of TiO2 NP

toxicity is primarily derived from studies of model

organisms, and these results have largely been

obtained from laboratory studies of cultures treated

with TiO2 NPs. Although such studies are useful for

validating the toxic effect of chemicals, they do not

consider environmental factors. Consequently, the

current understanding of the potential impact of TiO2

NPs on the environment is poor. Specifically, nitrify-

ing bacteria are sensitive to a number of environmental

conditions (i.e., pH, dissolved oxygen concentration,

and temperature) and are therefore susceptible to

inhibition (Zhen et al. 2011). Although silver NPs and

zinc oxide NPs are reportedly toxic to the respiration

processes of nitrifying bacteria (Choi and Hu 2008;

Zheng et al. 2011), little research exists on the

potential effect of TiO2 NPs on biological nitrogen

removal; therefore, this topic should be explored more

fully.

Nitrogen (N) is the key nutrient that causes

eutrophication in waterways. Therefore, nitrogen-

containing materials are compulsorily removed from

wastewater sources in most developed countries. An

important function of activated sludge is biological

nutrient removal (Zheng et al. 2011), which is carried

out via complex microbial populations that perform

nitrification and de-nitrification. Therefore, activated

sludge contains various species of bacteria, and it is

difficult to discern the potential impact of TiO2 NPs on

certain pure species of bacteria. Thus, the diversity of

the microbial populations and a stable bacterial

community structure play important roles in achieving

a high efficiency of biological nitrogen removal.

However, few research studies have been published on

the effect of TiO2 NPs on the bacterial community in

activated sludge. Although phosphorus (P) is also

assumed to be a key nutrient for the eutrophication of

freshwater, much like nitrogen (N), we found that

phosphorus removal was unaffected by the presence of

TiO2 NPs (Zhen et al. 2011). Therefore, we did not

investigate the effect of the presence of TiO2 NPs on

phosphorus removal.

The objectives of this study were to (a) determine

the effects of TiO2 NPs on biological nitrogen removal

from wastewater (b) determine the effect of TiO2 NPs

on the bacterial diversity of activated sludge, and

(c) determine whether natural light stimulates the

toxicity of NPs to the bacteria in activated sludge, thus

affecting biological nitrogen removal. In this study, a

sequencing batch reactor (SBR) known to achieve

nitrogen removal was used to culture the activated

sludge. A scanning electron microscope (SEM) was

used to assess the surface integrity of the activated

Biodegradation

123

sludge, and changes in the bacterial diversity of the

activated sludge were investigated using polymerase

chain reaction-denaturing gradient gel electrophoresis

(PCR-DGGE) analysis.

Materials and methods

Preparation of nanoparticle suspensions

The TiO2 NPs used in this study were purchased from

Sigma-Aldrich (St. Louis, MO, USA). The reported

mean particle size was less than 25 nm, and the purity

was 99.7 %, Different concentrations of TiO2 NPs in

aqueous solutions were prepared as previously

described (Keller et al. 2010). For example, to produce

a 100 mg L-1 NP stock suspension, 100 mg of TiO2

NPs was dispersed into 1 L of Milli-Q water via

sonication for 1 h (25 �C, 250 W, 40 kHz). The actual

size of the particles in suspension in water was

determined to be 16–34 nm using transmission elec-

tron microscopy (TEM) (Fig. 1).

Operation of SBRs

Three SBRs with working volumes of 5 L were seeded

with sludge from the Taiping sewage treatment plant

(Harbin, China) in a biological nutrient removal process

with intermittent (SBR-type) operation. All reactors

were fed with synthetic wastewater that contained the

following (mg L-1): NH4?–N (supplied from NH4Cl),

40; chemical oxygen demand (COD) (supplied from

sodium acetate), 300; KH2PO4, 110; NaHCO3, 500;

MgSO4�7H2O, 50; and CaCl2�2H2O, 10. A microele-

ment solution (1 mL L-1 of the wastewater) contained

(mg L-1): FeCl3�7H2O, 1.5; CuSO4�H2O, 0.03; EDTA,

10; MnCl2�4H2O, 0.12; KI, 0.18; ZnSO4�7H2O, 0.12;

CoCl2�7H2O, 0.15; and H3BO3, 0.15.

The laboratory scale SBR operated at three cycles

per day. Each cycle consisted of five phases, including

fill instant (5 min), aerobic phase (4 h), anoxic phase

(2 h), sludge settling (1 h), and effluent discharge

(5 min) in a temperature-controlled room (25–28 �C).

During the aerobic stage, air was provided intermit-

tently using an air pump controlled by an automatic

on/off switch according to an online DO detector to

maintain a DO level between 0.45 and 0.55 mg L-1.

The pH in the system was recorded but not

controlled, and it fluctuated between 7.4 and 8.2.

The sludge was removed to waste to maintain a solids

retention time (SRT) of approximately 22 days (Zhen

et al. 2011). The effluent concentrations of NH4?–N,

nitrite–nitrogen (NO2-–N), and nitrate–nitrogen

(NO3-–N) in all SBRs were frequently determined

until the nitrogen and phosphorus removals became

relatively stable (at approximately 90 days).

The effect of TiO2 NP accumulation on the

biological nitrogen removal from wastewater was

investigated by continuously dosing SBR 1 and SBR 2

with TiO2 NPs over a seven-day period, and SBR 3

was operated as the control (without the addition of

TiO2 NPs). The daily concentrations of TiO2 NPs in

SBR 1 and SBR2 could gradually decrease as a result

of effluent and sludge discharge; for this reason, TiO2

NPs were supplemented every day (during the first

cycle) to present final concentrations as follows:

0 mg L-1 (day 1), 2 mg L-1 (day 2), 10 mg L-1

(day 3), 25 mg L-1 (day 4), 50 mg L-1 (day 5),

100 mg L-1 (day 6), and 200 mg L-1 (day 7). This

supplementation was performed after determining the

total concentration of TiO2 in each reactor with the

goal of simulating the likely route of exposure in the

Fig. 1 TEM images of TiO2 NPs in suspension in water

Biodegradation

123

environment, because the environmental release of

TiO2 NPs might continuously increase due to their

large-scale production (Kiser et al. 2009).

SBR 1 was covered with aluminum foil to prevent

possible light-induced effects, and each SBR was set

up in triplicate. To examine whether natural light

stimulates the toxicity of TiO2 NPs, which may

subsequently affect biological nitrogen removal,

SBR2 was set up near an east-facing window in the

laboratory in September (N45.761, W126.688).

DNA extraction

The bacterial genomic DNA of the activated sludge

was extracted according to a previous publication

(Wan et al. 2011) using a DNA Isolation Kit (MO Bio

Laboratories, Inc., Carlsbad, CA, USA). The extracted

DNA was dissolved in 60 lL of TE buffer solution.

PCR-DGGE analysis of the bacterial community

in activated sludge

The V3 region of 16S rRNA was amplified by PCR

using the universal bacterial primers (338F, 5-ACT-

CCTACGGGAGGCAGCAG-3 and 534R, 50-ATTA-

CCGCGGCTGCTGG-3 with a GC clamp). The PCR

amplification was conducted in a 50-lL tube contain-

ing 5 lL of 109 PCR buffer (Mg2? Plus), 4 lL of

dNTP mixture (2.5 mM), 1 lL of 338F primer

(20 lM), 1 lL of 534R primer (20 lM), 2.5 ng of

DNA template, and 0.15 U of Taq DNA polymerase

(Takara, Dalian, China). The thermal cycler (model

9700; ABI, Foster, CA, USA) used a cycling program

that began with an initial denaturation of DNA for

10 min at 94 �C, followed by 30 cycles of 1 min at

94 �C, 30 s at 55 �C (decreasing by 0.1 �C per cycle to

52 �C), and 1 min 30 s at 72 �C. The final extension

step was performed for 10 min at 72 �C.

The PCR products were separated using the

DcodeTM universal mutation detection system (Biorad

Laboratories, Hercules, CA, USA). Polyacrylamide

gels with a 40–60 % vertical denaturing gradient

(100 % denaturant corresponds to 7 mol L-1 urea and

40 % deionized formamide) were prepared. A total of

10 lL of each PCR product was loaded, and the gels

were electrophoresed for 7.5 h at 150 V and 60 �C.

The gels were silver stained as described previously

(Bassam et al. 1991), scanned using a transmission

scanner, and examined using PHYLIP 4.0 and SPSS

12.0 software.

The dominant DGGE bands were excised and

dissolved in 20 lL of 19 TE. In total, 2 lL of DNA

solution was used as the PCR amplification template

under the conditions described above, and the same

primers were used. The PCR products were recovered

by agarose gel electrophoresis, ligated into the vector

pMD18 (Takara, Dalian, China), and cloned into

E. coli DH5a. Positive clones were examined for

ampicillin resistance via blue–white spot screening.

The positive clones were sequenced using an ABI3730

instrument, and partial 16S rRNA gene sequences

were analyzed using the BLAST program in GenBank.

The gene sequences of the clones were deposited in

GenBank under the accession numbers FJ440449

through FJ440511.

Scanning electron microscopy (SEM)

The SEM procedure used was as follows: after

exposure to TiO2 NPs, 5 mL of mixture sludge was

centrifuged at 4,000 rpm for 5 min, the supernatant

was discarded, and a pair of tweezers (or a syringe

needle) was used to extract the precipitate (a diameter

of 0.5 mm). The mud was placed in a 5 mL centrifuge

tube, and 2.5 % glutaraldehyde was added for 1.5 h at

4 �C. After rinsing three times with 0.1 M phosphate

buffer (pH 6.8), the mud was dehydrated in a series

of graded ethanol solutions (50, 70, 80, and 90 %,

15 min per step) and then dehydrated in 100 %

ethanol three times for 15 min each time. The ethanol

was then replaced 1:1 100 % ethanol:isoamyl acetate,

followed by pure isoamyl acetate, and the samples

were incubated each time for 15 min. The samples

were then placed in a desiccator and dried for 8 h.

A Giko IB-5 ion sputtering instrument was used to

spray gold at a thickness of 1,500 nm onto the sample

surface. The sample was viewed under a FEI Quanta

200 scanning electron microscope at 20 kV (FEI

Company, USA).

Transmission electron microscopy (TEM)

The TiO2 NP suspension was transferred onto a copper

grid. After the grid was dried in a laminar flow hood, it

was analyzed under a JEOL 2100 field emission gun

transmission electron microscope (Peabody, MA).

Biodegradation

123

Analytical methods

The determinations of NH4?–N, nitrite–nitrogen

(NO2-–N), and nitrate–nitrogen (NO3

-–N) were

conducted in accordance with Nessler’s reagent

spectrophotometry, N-(1-naphthalene)-diaminoethane

spectrophotometry, and ultraviolet spectrophotomet-

ric methods, respectively (APHA 1998). The dis-

solved oxygen (DO) concentration and the pH of the

suspensions were determined online using a WTW

Multi 340i DO meter and a WTW InoLab level 2

instrument, respectively.

Dehydrogenase activity assay

Dehydrogenase activity is often used as the activity

index for activated sludge and is measured by the

reduction of 2,3,5-triphenyltetrazolium chloride

(TTC). Two milliliters of Tris–Hcl buffer solution

(0.2 M, pH 7.3), 5 mL of TTC (5 g L-1), and 2 mL of

glucose solution (0.1 M) were added to 1 mL of

activated sludge. The mixture was shaken at 140 rpm

for 20 min at room temperature and was subsequently

incubated at 37 �C for 12 h. The deoxidization

reaction was stopped by the addition of 100 mL of

concentrated sulfuric acid. The follow-up steps were

performed as previously described (Lv et al. 2008).

Briefly, the absorbance of toluene was measured at

492 nm using a UV spectrometer. The enzyme activity

was expressed as the amount of TTC reduced per mg

of activated sludge per h. The dehydrogenase activity

was calculated using the following equation:

Edehydrogenase ¼ 0:08M= Vs�MLSSð Þ ð1Þ

where Edehydrogenase is the dehydrogenase activity

[lg (g MLSS)-1h-1], M is the TTC quantity calcu-

lated based on the standard line (lg), and Vs is the

volume of the activated sludge sample.

Results and discussion

Effects of TiO2 NPs on biological nitrogen

removal

Figure 2 shows the effluent concentrations of NH4?–N,

NO3-–N, and NO2

-–N in SBR1 over 7 days resulting

from TiO2 NP dosing concentrations of 0, 2, 10, 25, 50,

100, and 200 mg L-1. The effluent concentrations

of NH4?–N at TiO2 NP concentrations ranging from 0

to 200 mg L-1 were relatively stable over 7 days,

suggesting that these TiO2 NP concentrations did not

have a significant influence on NH4?–N removal. The

effluent concentrations of NO3-–N and NO2

-–N at

TiO2 NP concentrations ranging from 2 to 50 mg L-1

were also observed to be stable and similar to those

observed in the absence of TiO2 NPs (day 1). More than

78 % of the total nitrogen (TN) was removed from

SBR1 over the first 4 days. These results indicate that

TiO2 NP concentrations ranging from 2 to 50 mg L-1

had no adverse effects on nitrogen removal. However,

when activated sludge was exposed to 100 and

200 mg L-1 of TiO2 NPs, the respective effluent

NO3-N levels were 8.9 and 8.7 mg L-1, and the

effluent NO2-–N levels were 0.27 and 0.31 mg L-1,

respectively, which were significantly higher than those

observed in the absence of TiO2 NPs (NO3-–N:

4.7 mg L-1 and NO2-–N: 0.01 mg L-1). The corre-

sponding TN removal efficiencies were 36.5 % and

20.3 %, respectively, which were markedly lower than

those observed in the control test (80 %). These results

indicate that nitrogen removal was inhibited by higher

concentrations of TiO2 NPs (C100 mg L-1).

The effect of TiO2 NPs on the transformations of

NH4?–N, NO3

-–N, and NO2-–N was investigated. As

shown in Fig. 3, the variations of NH4?–N were not

significantly different among TiO2 NPs concentrations

ranging from 0 to 200 mg L-1. Additionally, the levels

of NO3-–N and NO2

-–N did not significantly differ in

Fig. 2 Effects of TiO2 NP concentration [0 mg L-1 (day 1),

2 mg L-1 (day 2), 10 mg L-1 (day 3), 25 mg L-1 (day 4),

50 mg L-1 (day 5), 100 mg L-1 (day 6), and 200 mg L-1 (day

7)] on the effluent concentrations of NH4?–N, NO3

-–N, and

NO2-–N. Error bars represent the standard deviations of

triplicate tests

Biodegradation

123

the presence of TiO2 NP concentrations of 0, 2, 10, 25,

and 50 mg L-1 (Fig. 4), indicating that the presence of

2–50 mg L-1 of TiO2 NPs had no acute effect on the

transformations of NO3-–N and NO2

-–N. However,

when the concentrations of TiO2 NPs reached 100 and

200 mg L-1, the NH4?–N removal efficiencies

remained high ([90 %), but the transformations of

NO3-–N and NO2

-–N were significantly inhibited.

Biological nitrogen removal is accomplished

through sequential nitrification, which involves the

oxidation of ammonium to nitrite and the oxidation of

nitrite to nitrate, in addition to de-nitrification pro-

cesses. This study showed that the decrease in

biological nitrogen removal induced by high concen-

trations of TiO2 NPs was not caused by the inhibition

of ammonia oxidation but was instead due to the

inhibition of de-nitrification. Similarly, a recent study

reported that high concentrations of other nanoparti-

cles, such as ZnO NPs (10 and 50 mg L-1), could

inhibit de-nitrification by the release of zinc ions

resulting from ZnO NP dissolution. However, in this

study, no titanium ions were detected in the presence

of 100 and 200 mg L-1 of TiO2 NPs. This result is in

agreement with those of Zhen et al. (2011) and Kiser

et al. (2009), who reported that in wastewater, Ti

occurs solely in the solid phase and not in ionic forms

due to its low solubility. A recent study suggested that

higher concentrations of TiO2 NPs could significantly

decrease TN removal by decreasing the abundance of

ammonia-oxidizing bacteria. However in this study,

we found that the decrease in biological nitrogen

removal induced by high concentrations of TiO2 NPs

was not caused by the inhibition of ammonia oxidation

but rather by the inhibition of de-nitrification. De-

nitrification involves the reduction of nitrate to nitrite,

Fig. 3 Effects of light and TiO2 NP concentration [0 mg L-1

(day 1), 2 mg L-1 (day 2), 10 mg L-1 (day 3), 25 mg L-1 (day

4), 50 mg L-1 (day 5), 100 mg L-1 (day 6), and 200 mg L-1

(day 7)] on the effluent concentrations of NH4?–N, NO3

-–N,

and NO2-–N. Error bars represent the standard deviations of

triplicate tests

Fig. 4 Variations in NH4?–N (a), NO3

-–N (b), and NO2-–N

(c) during one cycle of exposure to various concentrations of

TiO2 NPs. Error bars represent the standard deviations of

triplicate tests

Biodegradation

123

then the reduction of nitrite to nitric oxide (NO) and,

subsequently, to nitrous oxide (N2O) and molecular

nitrogen (N2), which are finally released into the

atmosphere. These transformations are carried out by a

group of bacteria that are capable of using nitrate in

place of oxygen as an electron acceptor for respiration

(Rodrıguez et al. 2011). In previous studies, TiO2 NPs

were revealed to be toxic to both Gram-negative and

Gram-positive bacteria (Adams et al. 2006), which

suggested that higher concentrations of TiO2 NPs in

the activated sludge might also decrease the abun-

dance of denitrifying bacteria. For this reason, the

effects of TiO2 NPs on changes in the bacterial

diversity of activated sludge were investigated.

Effects of light on biological nitrogen removal

We compared the biological nitrogen removal (work-

ing) efficiency between SBR1 and SBR2 (Figs. 2, 4)

and observed that the overall trends of the effluent

concentrations of NH4?–N, NO3

-–N, and NO2-–N

resulting from the addition of TiO2 NPs

(0–200 mg L-1) over a period of 7 days were similar

between the two reactors; this result suggests that the

inhibition of biological nitrogen removal occurred

under both dark and illuminated conditions, and that

illumination did not enhance the inhibition.

Previous studies have shown that the toxicity of TiO2

to bacteria under pure culture conditions was signifi-

cantly greater under light than dark conditions (Adams

et al. 2006), which supports the notion that the

antibacterial activity of TiO2 is related to photo-

catalytic ROS production. However, a significant

decrease in the abundance of ammonia-oxidizing

bacteria (AOB) in activated sludge that was exposed

to 50 mg L-1 of TiO2 under dark conditions still

occurred (Zhen et al. 2011), indicating that an undeter-

mined mechanism may be involved. In this study, the

inhibition of biological nitrogen removal by TiO2 was

similar under both dark and light conditions (Fig. 4),

which suggests that the toxicity of TiO2 to organisms

under pure culture conditions may not reflect natural

systems. The toxicity mechanisms of TiO2 NPs under

dark conditions should be investigated further.

Effects of TiO2 NPs on bacterial diversity

The mechanism of TiO2 NP-mediated de-nitrification

was explored by assessing the microbial community

via PCR-DGGE analysis. As shown in Fig. 5, the

bacterial diversity in activated sludge from SBR1

showed a clear decline over 7 days). This result

indicated that the bacterial abundance in activated

sludge was affected by the presence of TiO2 NPs. This

result was similar to those of previously published

reports showing that TiO2 NPs reduce the diversity of

the microbial community in activated sludge after

70 days of exposure (Zhen et al. 2011). Additionally,

Ge et al. (2011) found that TiO2 NPs were able to alter

the bacterial composition in soil and reduce microbial

populations after 60 days of exposure.

We identified 23 DGGE bands in the activated

sludge sample (Fig. 5), and these bands were excised

and sequenced using BLAST (Table 1). On the first

day, the dominant bacterial communities in the

activated sludge contained typical ammonia-oxidizing

(band 3, related to Nitrosomonas sp.) and denitrifying

bacteria (band 9, related to Pseudomonas sp.), and

these bacteria have been reported to remove nitrogen

from wastewater. However, after 7 days of exposure,

when the concentrations of TiO2 NPs reached

Fig. 5 DGGE profile of

bacterial communities in

SBR1. L1 and L 2 represent

activated sludge in SBR1

with the addition of

0 and 200 mg/L TiO2 NPs.

Detailed information on

bands 1–23 is presented in

Table 1

Biodegradation

123

200 mg L-1, the number of Pseudomonas sp. (band 9

in Fig. 5) declined in the activated sludge; this

microorganism is known to denitrify nitrate. The

above results may explain the inhibitory effect of

TiO2 NPs on the de-nitrification process. Conversely,

ammonia oxidizing bacteria were still observed in the

presence of 200 mg L-1 of TiO2 NPs in the activated

sludge. It is well known that nitrifying bacteria are

sensitive to environmental factors. Nevertheless, the

results of this study indicate that these microorganisms

may be able to tolerate high levels of TiO2 NPs.

Effects of TiO2 NPs on dehydrogenase activity

and the surface integrity of activated sludge

Dehydrogenase activity is a valuable indicator of

microbial viability in various environments, including

Table 1 DGGE bands and

their closely related

sequence

Band Result Accession number Similarity

(%)

L1 Roseomonas sp. AY624051.1 96

L2 Acidobacteria bacterium JF345309.1 99

L3 Uncultured Nitrosomonas sp.

Uncultured ammonia-oxidizing bacterium clone

FM997827.1 DQ154799.1 99

99

L4 Azospirillum sp. FJ386552.1 94

L5 Uncultured Dechloromonas sp. JF808881.1 95

L6 Uncultured Sphingobacterium sp. FN668067.2 99

L7 Uncultured bacterium clone HQ891402.1 99

L8 Uncultured Ruminococcaceae bacterium clone GQ358507.1 94

L9 Pseudomonas sp. HQ438076.1 99

L10 Sphingomonas sp. JF297632.1 96

L11 Uncultured Rheinheimera sp. GQ464399.1 100

L12 Uncultured Arenimonas sp. JN648282.1 94

L13 Uncultured Leptothrix sp. GU572372.1 99

L14 Uncultured Acidobacteria bacterium clone HQ598480.1 100

L15 Uncultured Zoogloea sp. EU639302.1 100

L16 Sphingomonas sp. JF716065.1 99

L17 Prevotella sp. HM245224.1 97

L18 Uncultured Dokdonella sp. JN679193.1 99

L19 Sphingomonas sp. JN848796.1 100

L20 Uncultured Rhodobacteraceae bacterium clone HQ003517.1 100

L21 Uncultured Dechloromonas sp. JN679130.1 97

L22 Acidobacteria bacterium JF345309.1 99

L23 Uncultured Sphingobacterium sp. FN668067.2 100

Fig. 6 Dehydrogenase activities of activated sludge in SBR1

and SBR3. SBR3 was a control reactor to which no TiO2 NPs

were added. Error bars represent the standard deviations of

triplicate tests

Biodegradation

123

activated sludge (Lv et al. 2008). The dehydrogenase

activity of an intracellular enzyme is affected by a

number of antimicrobial and bacteriostatic agents,

such as toluene and chloroform, which exist in various

environments. The effect of TiO2 NPs on dehydroge-

nase activity is presented in Fig. 6, which shows that

dehydrogenase activity was inhibited by all of the

treatments containing TiO2 NPs. Figure 5 shows that

the activity of dehydrogenase decreased with

increased concentrations of TiO2 NPs. Different

concentrations had significant effects on dehydroge-

nase activity, which were likely due to decreased

microbial activity. SEM analysis also revealed the

change of microorganism’s composition structure and

damage on the surface of activated sludge that was

exposed to 100–200 mg L-1 TiO2 NPs (Fig. 7). These

results revealed that high concentrations of TiO2 NPs

may disrupts the balance of microbial populations,

which may be responsible for the antibacterial activity

of TiO2 NPs.

Conclusions

In this study, the potential effects of TiO2 NP

accumulation on biological nitrogen removal from

wastewater and bacterial community shifts in acti-

vated sludge were investigated by continuous dosing

of SBRs with TiO2 NPs. Although TiO2 NP concen-

trations ranging from 2 to 50 mg L-1 had no adverse

effects on nitrogen removal, nitrogen removal was

dramatically inhibited by TiO2 NP concentrations of

100–200 mg L-1, possibly due to the inhibition of de-

nitrification. It was observed that 100–200 mg L-1 of

TiO2 NPs reduced the microbial diversity in activated

sludge, and the abundance of denitrifying bacteria was

significantly decreased, possibly explaining the inhib-

itory effect. Further studies showed that the activity of

dehydrogenase decreased with increased TiO2 NP

concentrations, and high concentrations of TiO2 NPs

damaged the activated sludge surface, thus proving

that high concentrations of TiO2 NPs possess antibac-

terial activity. Moreover, we also examined whether

Fig. 7 SEM images of activated sludge in SBR1. a–c represent

the activated sludge in SBR1 (without the addition of TiO2 NPs,

with the addition of 100 mg L-1 TiO2 NPs, and with the

addition of 200 mg L-1 TiO2 NPs, respectively)

b

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123

natural light stimulates the toxicity of TiO2 NPs,

subsequently affecting biological nitrogen removal.

The results showed that the inhibition of biological

nitrogen removal by TiO2 NPs was similar under both

dark and light conditions, suggesting that the toxicity

mechanisms of TiO2 NPs under dark conditions

require further investigation.

Acknowledgments This work was supported by the National

Natural Science Foundation of China (Grant No. 51078102) and

the Funds for Creative Research Groups of China (Grant No.

51121062).

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