Bioresource Technology 134 (2013) 180–189

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Bioresource Technology

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Long term assessment of factors affecting nitrifying bacteriacommunities and N-removal in a full-scale biological processtreating high strength hazardous wastewater

0960-8524/$ - see front matter � 2013 Elsevier Ltd. All rights reserved.http://dx.doi.org/10.1016/j.biortech.2013.02.036

⇑ Corresponding authors. Tel.: +82 51 200 7620; fax: +82 51 208 8190 (Y.M. Kim),tel.: +82 54 279 2275; fax: +82 54 279 2699 (J.M. Park).

E-mail addresses: kim78051@postech.ac.kr (Y.M. Kim), jmpark@postech.ac.kr(J.M. Park).

1 These authors contributed equally to this work.

Young Mo Kim a,⇑,1, Hongkeun Park b,1, Kyung Hwa Cho c, Jong Moon Park d,⇑a Department of Civil Engineering, Dong-A University, Busan 604-714, Republic of Koreab Department of Earth and Environmental Engineering, Columbia University, New York, NY 10027, USAc School of Urban and Environmental Engineering, Ulsan National Institute of Science and Technology, Ulsan 689-798, Republic of Koread Department of Chemical Engineering, School of Environmental Science and Engineering, Pohang University of Science and Technology, Pohang 790-784, Republic of Korea

h i g h l i g h t s

" We characterized the dynamics ofnitrifying bacterial communities in afull-scale WWTP.

" Full nitrification led to goodperformance in TN removal and ahigher nitrification rate.

" Amounts of AOB were positivelylinked to influent SCN�

concentration." Nitrobacter and nitrite

concentrations were closelyassociated.

" COD and phenol concentrationsseemed the primary reason fordeteriorated nitrification.

g r a p h i c a l a b s t r a c t

a r t i c l e i n f o

Article history:Received 1 December 2012Received in revised form 12 February 2013Accepted 14 February 2013Available online 26 February 2013

Keywords:Activated sludgeFull-scale processNitrifying bacteriaNitrificationPCA

a b s t r a c t

Over a 3 year period, interactions between nitrifying bacterial communities and the operational param-eters of a full-scale wastewater treatment plant were analyzed to assess their impact on nitrification per-formance. Throughout the study period, nitrification fluctuated while Nitrosomonas europaea andNitrosomonas nitrosa, the two major ammonia oxidizing bacteria (AOB) communities, showed resistanceto changes in operational and environmental conditions. Nitrobacter populations mostly exceeded thoseof Nitrospira within nitrite oxidizing bacteria (NOB). Meanwhile, principal component analysis (PCA)results revealed that a close association between Nitrobacter and nitrite concentration as well as a directcorrelation between the quantity of AOB and influent SCN� concentration. The serial shifts of data pointsover time showed that the nitrification of a full-scale treatment plant has been gradually suppressed bythe influence of influent COD and phenol concentrations.

� 2013 Elsevier Ltd. All rights reserved.

1. Introduction

Biological nitrogen removal (BNR) in wastewater treatmentplants (WWTPs) requires the oxidation of ammonia to nitratethrough nitrification and the reduction of nitrate to nitrogen gasthrough denitrification, thereby removing the nitrogen from thewastewater. Of these, aerobic nitrification plays a critical role in

Y.M. Kim et al. / Bioresource Technology 134 (2013) 180–189 181

BNR in an activated sludge system. This nitrification is performedin two consecutive steps by chemolithoautotrophic bacteria.Aerobic ammonia oxidizing bacteria (AOB) oxidize ammonia-N tonitrite-N with hydroxylamine as an intermediate product—theso-called nitritation process. The nitrite-N is subsequently oxidizedto nitrate-N in a single step by aerobic nitrite oxidizing bacteria(NOB)—a process denoted as nitratation. Aerobic AOB that prolifer-ate during the biological treatment of wastewater are primarilyb-proteobacteria and include the Nitrosomonas and Nitrosospira lin-eages. The NOB are a diverse group containing Nitrobacter andNitrospira (Juretschko et al., 2002; Purkhold et al., 2000).

In engineered BNR systems, nitrification is the rate-limiting stepand its recovery requires a long time once nitrifying bacteria havebeen washed out of the WWTP because of low specific growthrates and biomass yields. In addition, the activity of nitrifying bac-teria in WWTPs can be adversely affected by shifts in the process’spH, temperature, dissolved oxygen levels, retention time, aeration/mixing intensities and the presence of toxic compounds (Kim et al.,2011a). The AOB population was significantly affected by the tem-perature in a bioreactor treating saline wastewater (Park et al.,2009). Inconsistent nitrification was attributed to diminishedactivities of nitrifying bacteria as a result of toxic inhibition bythe wastewater (Kim et al., 2011a). The failure of ammonia oxida-tion occurred because of the low population of AOB (Wanner et al.,2004). It has also been reported that unstable nitrification perfor-mance was due to AOB–NOB mutualism (Graham et al., 2007).

Many of these critical process failures arising during biologicalnitrogen removal are likely attributable to variations in the relativeamount or activity of the AOB or NOB communities. Such varia-tions are thought to result from a combination of various proper-ties such as reactor design and environmental factors (Wellset al., 2011). Therefore, understanding and predicting what envi-ronmental conditions determine the dynamics of a nitrifying bac-teria community and its function is vital to successfully achievebiological nitrogen removal in WWTPs.

However, many of these studies have occurred in lab-scale reac-tors, where selective pressures may dramatically differ from thosein full-scale plants (Seviour and Nielsen, 2010). Even more, thecauses for nitrification breakdown events in full-scale systemsare not always obvious. Also, there have been few studies to datewhich have explored full-scale systems treating high strengthindustrial wastewater containing high levels of toxic N- andC-compounds like thiocyanate, cyanide and phenol (Table 1)(Figuerola and Erijman, 2010; Kim et al., 2011a).

The primary objectives of this study were to monitor the longterm dynamics in a nitrifying bacteria community and N-removalof an activated sludge system treating high strength hazardouswastewater as well as to identify those environmental factors thatmost significantly correlate with these dynamics with the goal ofimproving the efficiency and functional stability of the systems.Accordingly, samples were collected every 2 months, and 16 envi-ronmental factors were monitored concurrently over 3 years at afull-scale activated sludge process treating wastewater from a coke

Table 1The composition of wastewater.

Analyses Value

COD 1824–4592 mg/LPhenol 154–760 mg/LSCN� 180–859 mg/LTotal cyanide (TCN) 6.2–45.8 mg/LTotal nitrogen (TN) 168–371 mg-N/LNH4

+ 34–167 mg-N/LNO2

� <0.1 mg-N/LNO3

� <0.1 mg-N/LOil 0.4–30 mg/L

plant. These samples were analyzed via 16S rRNA gene-based Ter-minal Restriction Fragment Length (T-RFLP) to identify the AOBlineages, and quantitative PCR (qPCR) to measure the quantitiesof both the AOB and NOB. Principal component analysis (PCA)was employed to identify those environmental variables most sig-nificantly associated with these dynamics.

2. Methods

2.1. Wastewater treatment plant operation and samples

A full-scale WWTP of a coke manufacturing plant in Kwangy-ang, South Korea employs a single sludge along with the recyclingof the nitrified effluent—the pre-denitrification activated sludgeprocess (Fig. 1). This system is composed of an anoxic tank(1650 m3) and an aerobic tank (4185 m3) with a combined volumeof 5835 m3 and daily treats approximately 1440 m3 of coke plantwastewater. Nitrified effluent was recycled from the aerobic tankto the anoxic tank at a rate of about 180 m3 per h (Fig. 1). Returnactivated sludge collected from the secondary clarifier waspumped back to the anoxic tank at a rate of about 80 m3 per h.The aerobic tank is aerated at a rate of about 4150 m3 per h. Inthe course of this study, the concentrations of pollutants in theraw wastewater varied as can be seen in Table 1.

During this study, the mixed-liquor suspended solids (MLSS) ofthe system were controlled in the range of 2200–3300 mg/L; theaverage hydraulic retention time (HRT) was 4.0 days; and the aver-age solid retention time (SRT) was 25 days. The SRT was controlledby removal of excess sludge, resulting in different MLSS concentra-tions in the system. The temperature range of each tank varied be-tween 29.4 and 35.6 �C. The average pH of the influent, anoxic tankand aerobic tank was maintained at 9.0, 7.8 and 7.0, respectively.The dissolved oxygen (DO) concentration in the aerobic tank wasmaintained in the range between 2.0 and 5.0 mg/L, while the DOlevel of the anoxic tank was maintained below 0.1 mg/L. Functionalstability of the system was defined and quantified by the effluentconcentration of total nitrogen (TN). MLSS samples for this studywere taken directly from the aerobic tanks every 2 months fromJanuary 2007 to November 2009. Nitrification activity was mea-sured directly on fresh samples. For the DNA based studies, eachsample of 1.0 mL was dispensed into a 1.5 mL sterile tube and cen-trifuged at 13000 g for 10 min. The supernatant was decanted andthe pellet was then stored at �80�C before being used.

2.2. Chemical analysis

The collected samples were centrifuged at 3500 rpm for 3 min(MF550, Hanil Sci. Ind., Korea), and the supernatants were ana-lyzed as follows: according to standard methods, chemical oxygendemand (COD), ammonia, phenol and thiocyanate (SCN�) wereanalyzed by the colorimetric method with a spectrophotometer(Genesys TM-5, Spectronic Inc., USA). After distillation, the cyanide(CN�) concentration was determined by the pyridine-pyrazolonemethod. Nitrite and nitrate ions were measured with an ion chro-matograph (ICS-1000, Dionex Co., USA). Total organic carbon(TOC), inorganic carbon (IC), and TN were measured with a TOC/TN analyzer (TOC-V csu, TNM-1, Shimadzu Co., Japan).

2.3. Microbial activity test

To investigate microbial activities of nitrifying bacteria in theaerobic tank at the WWTP, the potential nitrification rate was esti-mated every 2 months for 3 years (January 2007–November 2009)through batch experiments. Batch experiments for nitrificationactivity were carried out in 500 mL Erlenmeyer flasks filled with

Fig. 1. Schematic diagram of a full-scale activated sludge process.

182 Y.M. Kim et al. / Bioresource Technology 134 (2013) 180–189

100 mL of test solution containing 50 mg-N/L of ammonium, 1.0 g/L NaHCO3, buffer solution (K2HPO4 2.5 g/L and KH2PO4 0.5 g/L) and1 mL trace metals/L (per liter: MgSO4 120.4 mg/L, CaCl2 111.0 mg/L, ZnSO4�7H2O 287.5 mg/L, CoCl2�6H2O 237.8 mg/L, FeSO4�7H2O278.0 mg/L, MoO3 200.0 mg/L). Without any pretreatment eachflask was inoculated with fresh activated sludge, obtained directlyfrom the full-scale aerobic tank, and then agitated with an excessof oxygen on a thermostatic shaker at 200 rpm and 35 �C, main-taining the pH at 7.5 by adding 1 M NaOH solution. The specificnitrification rate was calculated applying the equation providedby Kim et al. (2008).

2.4. DNA extraction, T-RFLP analysis, cloning and sequencing

All DNA in the pellet was extracted using an automated nucleicacid extractor (Magtration System 6GC, PSS, Chiba, Japan). PurifiedDNA was eluted with 100 lL of Tris–HCl buffer (pH 8.0) and storedat �20 �C for further analyses.

Temporal dynamics of AOB lineages were monitored via T-RFLPusing the protocol of a previous study based on the known 16SrRNA gene of AOB (Siripong and Rittmann, 2007). Because of thelow concentration of DNA from the nitrifiers, we amplified theDNA through nested PCR, using the universal primers 11f and1492r (Table 2), followed by the specific amplification of the nitri-fier gene (Nitrifier-specific reverse primer: Nso1225r, Forward pri-mer: Eub338f included phosphoramidite dye 6-FAM (Table 2))(Siripong and Rittmann, 2007). Digested PCR products with MspIrestriction endonuclease were run through an ABI 3130XL GeneticAnalyzer (Applied Biosystems, Foster City, USA) at the SolGentCompany (Korea). The peak results were analyzed using the Peak

Table 2Primers and probes used in T-RFLP and qPCR.

Target Primer/probe Sequence (5’–3’)

For T-RFLPBacterial 16S rDNA 11f 5’-GTTTGATCCTGGCTCA

1492r 5’-TACCTTGTTACGACTT-Bacterial 16S rDNA Eub 338f 5’-(6-FAM)-ACTCCTACGGAOB 16S rDNA Nso 1225r 5’-CGCCATTGTATTACGT

For qPCRBacterial 16S rDNA 1055f 5’-ATGGCTGTCGTCAGCT

1392r 5’-ACGGGCGGTGTGTAC-16STaq1115 5’-(6-FAM)-CAACGAGCG

AOB 16S rDNA CTO 189fA/Ba 5’-GGAGRAAAGCAGGGGCTO 189fCa 5’-GGAGGAAAGTAGGGGRT1r 5’-CGTCCTCTCAGACCARTMP1 5’-(6-FAM)-CAACTAGCT

Nitrospira spp. 16S rDNA NSR 1113f 5’-CCTGCTTTCAGTTGCTANSR 1264r 5’-GTTTGCAGCGCTTTGTNSR 1143Taq 5’-(6-FAM)-AGCACTCTG

Nitrobacter spp. 16S rDNA Nitro 1198f 5’-ACCCCTAGCAAATCTCNitro 1423r 5’-CTTCACCCCAGTCGCTGNitro 1374Taq 5’-(6-FAM)-AACCCGCAA

a A mixture of CTO 189fA/B and CTO 189fC at the weight ratio of 2:1 was used as the

Scanner software v 1.0 (https://products.appliedbiosystems.com;Applied Biosystems, Foster City, USA). Details of the PCR condi-tions, product purification and restriction digestion are providedelsewhere (Siripong and Rittmann, 2007). The relative amountsof each terminal fragment (TF) were calculated as the ratio of thepeak height for that TF to the sum of the peak height for all TFsin a given profile and expressed as a percentage.

Prior to cloning, the amplified unlabelled AOB 16S rRNA genefragments were purified using the PCR purification kit (SolGent,Korea). Purified PCR products were ligated into pGEM-T Easy clon-ing vectors (Promega, USA) and transformed into competent E. coliwith One-Shot Mach 1-T1 (Invitrogen, USA), as described in themanufacturer’s protocol. About fifty white colonies were selectedand cultivated. Primers T7 and SP6 were used to perform colonyPCR and to verify that the insert size was correct. Following PCRconfirmation of the insert size, amplified inserts were run on2% (wt/vol) agarose gels. The samples containing inserts of the esti-mated size were used for subsequent sequencing. The 16S rRNAgene inserts were sequenced through an ABI 3130XL Genetic Ana-lyzer (Applied Biosystems, Foster City, USA) at the SolGent Com-pany (Korea). Database homology searches for these sequenceswere performed using the BLAST program in the National Centerfor Biotechnology Information (NCBI) database.

2.5. qPCR analysis

To determine the quantities of the nitrifying bacteria, four inde-pendent qPCR assays were conducted in triplicate using the Taq-Man probes targeting total bacterial 16S rRNA gene, AOB 16SrRNA gene, Nitrospira spp. 16S rRNA gene, and Nitrobacter spp.

References

G-3’ Kane et al. (1993)3’ Lin and Stahl (1995)GAGGCAGC-3’ Amann et al. (1990)

GTGA-3’ Mobarry et al. (1996)

-3’ Ferris et al. (1996)3’ Ferris et al. (1996)CAACCC-(TAMRA)-3’ Harms et al. (2003)ATCG-3’ Hermansson and Lindgren (2001)ATCG-3’ Hermansson and Lindgren (2001)

CTACTG-3’ Hermansson and Lindgren (2001)AATCAGRCATCRGCCGCT-(TAMRA)-3’ Hermansson and Lindgren (2001)

CCG-3’ Dionisi et al. (2002)ACCG-3’ Dionisi et al. (2002)AAAGGACTGCCCAGG-(TAMRA)-3’ Harms et al. (2003)AAAAAACCG-3’ Graham et al. (2007)

ACC-3’ Graham et al. (2007)GGAGGCAGCCGACC-(TAMRA)-3’ Graham et al. (2007)

forward primer.

Y.M. Kim et al. / Bioresource Technology 134 (2013) 180–189 183

16S rRNA gene (Table 2). Each capillary tube was separately loadedwith 2 lL of template, followed by 4.0 pmol of the forward and re-verse primers (1 lL), together with 2.0 pmol of the TaqMan probe(0.5 lL) corresponding to each primer and probe set, 12.5 lL ofTaqMan Universal PCR Master Mix (No. 4304437 Applied Biosys-tems, New Jersey, USA), and PCR-grade sterile water for a final vol-ume of 25 lL. The amount of total bacterial 16S rRNA gene wasamplified using primer (1055f and 1392r) (Ferris et al., 1996).The TaqMan probe (16STaq1115) was modified by the 1114f pri-mer (Harms et al., 2003). The PCR program was 2 min at 50 �C,10 min at 95 �C; 45 cycles of 30 s at 95 �C, 60 s at 50 �C, and 40 sat 72 �C. To determine the amount of AOB 16S rRNA gene, two for-ward primers (CTO 189A/B and CTO 189C), one reverse (RT1r), andthe TaqMan probe (TMP1) were used as described previously byHermansson and Lindgren (2001). The PCR program for AOB 16SrRNA gene quantification included 2 min at 50 �C, 10 min at95 �C; 40 cycles of 30 s at 95 �C, 60 s at 60 �C. The Nitrospira spp.16S rRNA gene primers (NSR 1113f/NSR 1264r) (Dionisi et al.,2002) and the TaqMan probe (NSR 1143Taq) (Harms et al., 2003)were tested. PCR amplification consisted of 2 min at 50 �C,10 min at 95 �C; 50 cycles of 30 s at 95 �C, 60 s at 60 �C. Lastly,the amount of Nitrobacter spp. from Graham et al. (2007) wasamplified using primer (Nitro 1198f/Nitro 1423r) and TaqManprobe (Nitro 1374Taq). The conditions used for amplification were2 min at 50 �C, 10 min at 95 �C; 50 cycles of 20 s at 94 �C, 60 s at58 �C, and 40 s at 72 �C. Standards were prepared using serially10-fold diluted plasmid DNA where the target genes were inserted.Standard curves for total bacteria, AOB, Nitrospira spp. and Nitrobact-er spp., quantification were generated by plotting the threshold cyclevalues versus log10 of the gene copy numbers. The amplification effi-ciency (E) was estimated using the slope of the standard curvethrough the following formula: E = (10�1/slope) � 1. The efficiencyof PCR amplification for each gene was between 90% and 100%.

2.6. Principal component analysis (PCA)

Principal component analysis (PCA) was applied to identify anystatistically significant relationship between influent/effluentwater qualities and microbial activity in this study. PCA is usedto reduce the dimension of a dataset consisting of a large numberof variables, while the variation of a dataset is presented in the ori-ginal data set (Jolliffe, 2002). Orthogonal transformation on the ori-ginal variables is used to obtain a new set of variables which isdesignated a principal component (PC). A non-parametric Kol-mogorov–Smirnov test was applied to verify the normality of thedataset and demonstrated that all variables followed normal distri-butions. If a variable is not followed by a normal distribution, thelogarithmic transformation is applied to convert it. Principal com-ponents whose eigenvalues are greater than 1 were extracted tointerpret the relationships in the dataset. The VARIMAX rotationwas then applied to obtain an easily interpretable rotated matrixbecause doing so allows even distribution of the variables in thematrix. Finally, PC scores were calculated by using the original dataand the rotated PCs matrix.

3. Results and discussion

3.1. Variations in process performance and operational conditions

The performance of a full-scale WWTP with its fluctuations inoperational and environmental variables over a 3 year period isshown in Fig. 2 and Table 1. It was hypothesized that fluctuationsin these variables may be significant drivers of N removal and thenitrifying bacterial community. The average influent temperaturewas 32 ± 4.0 �C with pH being 9.0 ± 1.0 for the period (Fig. 2(a)and (b)). Influent temperature is relatively higher than that of

ordinary sewage since the wastewater flowing into the main sys-tem is generated from hot steam processes such as coal coking,coal gas purification and the by-product recovery process of thecoke plant. To prevent deterioration in the process performanceby the higher temperature of activated sludge, the influent temper-ature maintained at below 35 �C through cool water especially dur-ing the summer seasons. The pH of the influent is not allowed toexceed 10.0 by adding H3PO4 solution in the equalization. AddedH3PO4 solution serves to control the pH of the influent while pro-viding a nutrient source for microorganisms in the activatedsludge. The inflow rate of coke plant wastewater into the systemranged from 40 to 60 m3 per h and the internal recycling ratio var-ied between 2.7 and 6.3 (Fig. 2(c) and (d)).

During the study phenol flowed into the full-scale process in therange of 154–833 mg/L. Most of it was first degraded in the anoxictank with any remaining phenol being almost completely degradedin the aerobic tank (Fig. 2(e)). COD in the range of 1824–5104 mg/Lwas fed into the anoxic tank where about 70% of it was subse-quently removed through biodegradation and dilution via internalrecycling. Residual COD flowed into the aerobic tank and was thenbiodegraded, resulting in a COD removal efficiency of over 80%(Fig. 2(f)). Almost all SCN�, which varied from 223 to 1040 mg/Lin the influent, was completely removed in the aerobic tank(Fig. 2(g)). SCN� is aerobically degraded to a combination ofammonia, carbonate, and sulfate by several species of chemolitho-trophic bacteria, Thiobacillus sp., which uses SCN� as a carbon andnitrogen source (Kim et al., 2011b). Total cyanides in the range of10.6–45.8 mg/L flowed into the system and 1.5–14.3 mg/L of itwas outflowed (Fig. 2(h)). The incomplete removal of total cya-nides resulted from the existence of ferric cyanide (Fe(CN)6

3–),which is known to undergo very slow biodegradation (Kim et al.,2008). To achieve the legal discharge level (1.0 mg/L in Korea) forthe residual total cyanides, a post-chemical precipitation processusing a ferrous sulfate solution and other cohesive agents has beenutilized in this facility.

3.2. Nitrification performance of the plant

Ammonia removal, nitrite/nitrate production and effluent TNconcentration fluctuated widely in the full-scale activated sludgeprocess (Fig. 3(a)–(c)). During steady state of full-nitrification (Jan-uary 2007, August 2007, December 2007, January 2008, May 2008,December 2008), the final effluent concentration of NH4

+-N in theaerobic tank remained less than 30 mg-N/L, while the NO3

� con-centration was in the range of 150–300 mg/L (about 34–68 mg-N/L). Detected NO2

� concentration values were less than 5.0 mg/L. The average TN in the influent was 250 mg-N/L. The effluentTN concentration remained below 50 mg-N/L, corresponding toan average TN removal efficiency of 80% (note that regulationsstipulate a TN concentration less than 60 mg-N/L for discharge intosurface water in South Korea). Also, the effluent TN concentrationof less than 50 mg-N/L indicated that the system was functionallystable throughout the study.

Concurrently, nitrification performance of the full-scale WWTPtreating high strength pollutants targeted in our study was oftenunstable during the 3 year period. Nitrification inhibition resultedin accumulation of ammonia and occasionally of nitrite. Periodsof low nitrification coincided with an increase in the effluent TNconcentration of over 100 mg-N/L (Fig. 3(a)–(c)), exceeding the le-gal discharge limit for TN concentration. Once nitrification failureoccurred in the WWTP, recovery of the nitrification process re-quired very long times (several months; 2–5 months) owing tothe slow growth rates of nitrifying bacteria. Although poor nitrifi-cation performance led to the collapse of TN removal efficiency ofthe full-scale WWTP, it did not decrease the efficient removal ofcarbon and other pollutants.

Fig. 2. Variation in (a) influent temperature; (b) influent pH; (c) inflow rate; (d) internal recycling ratio; influent- and effluent-concentrations (e) phenol; (f) COD; (g) SCN�;(h) TCN in the full-scale process during the study period (solid circles: influent, solid triangles: effluent).

Tota

l Nitr

ogen

(mg-

N/L

)

0

100

200

300

400

Operation date

01/01

/2007

07/01

/2007

01/01

/2008

07/01

/2008

01/01

/2009

07/01

/2009

01/01

/2010

Nitr

ite, N

itrat

e (mg/

L)

050

100150200250300350

(a)

(b)

(c)

Am

mon

ium

(mg-

N/L

)

020406080

100120140 (d)

Time

Nitr

ite, N

itrat

e (mg-

N/L

)

0

10

20

30

40

50 (e)

01/29

/2007

03/26

/2007

06/20

/2007

8/28/2

007

10/22

/2007

12/27

/2007

01/21

/2008

03/28

/2008

05/16

/2008

08/05

/2008

10/17

/2008

12/10

/2008

01/07

/2009

04/25

/2009

06/25

/2009

08/05

/2009

09/28

/2009

11/20

/2009

Spec

ific

nitr

ifica

tion

rate

(m

g-N

/ g-

VSS

h)

012345678 (f)

Sampling date

Am

mon

ium

(mg-

N/L

)

050

100150200250300350 (b)

Effluent NO2-

Effluent NO3-

Effluent NO2-

Effluent NO3-

(A) (B)

Fig. 3. (A) During the whole 3 year period, variation in (a) TN; (b) ammonia; (c) nitrite and nitrate concentrations; (B) At points sampled bimonthly for 3 year period,variation in (d) ammonia; (e) nitrite and nitrate concentrations (f) specific nitrification rate in the aerobic tank (solid circles: influent, solid triangles: effluent).

184 Y.M. Kim et al. / Bioresource Technology 134 (2013) 180–189

Bimonthly for the 3 year study period, variations in nitrificationactivity were monitored thorough batch experiments using freshlyactivated sludge sampled directly from the full-scale aerobic tank.As illustrated in Fig. 3(d) and (e), nitrifying bacteria activity (poten-tial nitrification rate) changed dynamically along with ammoniaremoval performance of the full-scale WWTP during the same per-iod. Generally, during periods of the lowest effluent ammonia con-

centrations (below about 30 mg-N/L) in the full-scale WWTP, thespecific nitrification rate attained the highest levels in the rangeof about 5.0–6.0 mg-N/g-VSS h. On the other hand, when highereffluent ammonia concentration (over about 100 mg-N/L) was dis-charged from the full-scale WWTP, the nitrifying bacteria activityseverely decreased, falling below the level of 1.0 mg-N/g-VSS h(Fig. 3(d) and (f)).

Fig. 4. Relative quantities (%) of each AOB species determined by the corresponding normalized T-RFLP profile peak area for the three species.

Table 3Expected TF sizes and their corresponding AOB group based on T-RFLP of 16S rRNAgene (Siripong and Rittmann, 2007).

TF size (bp) Nitrifying bacteria group

164–166, 276 Nitrosomonas europaea/eutropha lineage276 Nitrosomonas oligotropha lineage276 Nitrosomonas cryotolerans lineage166 Nitrosomonas marina lineage276 Nitrosomonas communis lineage105–107 Nitrosospira lineage

Y.M. Kim et al. / Bioresource Technology 134 (2013) 180–189 185

The pattern of effluent nitrate levels in the aerobic tank was sim-ilar to variations in nitrification activity (Fig. 3(e) and (f)). In re-sponse to a gradual decrease in nitrate concentrations (from 47 to17.5 mg-N/L) in January to June 2007, the batch test also revealeda decrease in nitrification activity from 4.75 to 2.57 mg-N/g-VSS h.At the lowest nitrite and nitrate levels (below 1.0 mg-N/L) in the fi-nal effluent, the nitrification rates in the activated sludge were thelowest, corresponding to below 1.0 mg-N/g-VSS h. After the in-creased nitrate concentrations in December 2007, May 2008 andDecember 2008, the nitrification rate attained maximum levels.Based on the relationships among the variations in TN removal, ni-trate concentrations and nitrification rates, the full conversion fromammonia into nitrate by nitrifying bacteria led to improved perfor-mance of TN removal and higher nitrification rate in the full-scaleWWTP. Consequently, high activities of both AOB and NOB were re-quired to improve the operational stability of this process treatinghigh strength hazardous wastewater through full nitrification.

Although changes in nitrifying bacterial activity (potential nitrifi-cation rate) usually correlated with the overall nitrification perfor-mance of the full-scale process well, there was discordancebetween the potential rate (batch test) and the full-scale perfor-mance in August 2008 and September 2009. One may conclude that,in spite of the potentially beneficial activity of nitrifying bacteria,unidentified environmental factors influenced their actual perfor-mance in the full-scale process, leading to nitrification failure.

3.3. AOB community structure analyzed with T-RFLP

T-RFLP of the 16S rRNA gene for AOB was conducted for a 3 yearsampling period. Fig. 4 illustrates the dynamics for the three fre-quently detected AOB lineages in the aerobic tank of the full-scale

process. As described in the materials and methods section, the rel-ative quantities of the three AOB groups were inferred from thenormalized TF peak areas corresponding to each AOB group. In typ-ical T-RFLP electropherograms for the AOB lineages (data notshown), TFs detected during our series were 102, 164 and276 bp. Of these, the 164 bp TF peak displayed the highest normal-ized peak areas for all samples although its relative peak areas fluc-tuated over time. In previous studies (Siripong and Rittmann,2007; Kim et al., 2011a), the peak at 164 bp was most frequentlyassociated with the Nitrosomonas europaea/eutropha lineage (Ta-ble 3). Besides the major peak at 164 bp, a peak at 276 bp was alsodetected, potentially representing the presence of N. europaea/eutropha, N. oligotropha, N. cryotolerans, or N. communis lineage (Ta-ble 3). A small peak at 102 bp, indicating the presence of the Nitr-osospira lineage, was also observed in samples from this WWTP.

To obtain more detailed understanding of the AOB communitypresent in the process, AOB 16S rRNA gene-based cloning andsequencing was performed showing that AOB clones were closelyassociated with N. europaea in the N. europaea/eutropha lineage,N. nitrosa in the N. communis lineage and the Nitrosospira lineage.

186 Y.M. Kim et al. / Bioresource Technology 134 (2013) 180–189

Consequently, the expected AOB groups corresponding to thepeaks at 102, 164, and 276 bp could be identified as Nitrosospira,N. europaea, and N. nitrosa lineage, respectively (Table 3). Inferredfrom the highest peak areas in the range of about 70–90%, N. euro-paea within AOB was most dominant in this wastewater treatmentsystem, irrespective of variations in environmental conditions orperformance of the full-scale process. N. nitrosa constituted asum total of about 20% of the AOB population in the samples andnever exceeded 40% of the AOB population. Thus, N. europaea andN. nitrosa constituted the two major AOB lineages in this studyand their dominance was not dynamically changed despite fluctu-ations in their relative quantities. Nitrosospira maintained evenlower levels than the other two major lineages, continuously rep-resenting less than 3% of the AOB population. Indeed, a singleAOB lineage, N. europaea dominated in this full-scale systemthroughout the entire time period.

The N. europaea lineage is known to be able to function ecolog-ically as both r-strategists (high specific growth rates and low sub-strate affinity) and K-strategists (low specific growth rates andhigh substrate affinity) (Ahn et al., 2008) reflecting the dominantpresence of the N. europaea lineage in WWTPs with a wider rangeof ammonia concentrations. (Ahn et al., 2008; Figuerola and Erij-man, 2010; Lydmark et al., 2007). Meanwhile, N. nitrosa, which cor-responds to the N. communis lineage, has previously been detectedon occasion in industrial WWTPs (Kim et al., 2011a,b; Layton et al.,2005). Their relatively high ammonia Ks value in the range of19–46 lM (Koops and Pommerening-Röser, 2001) might explaina better adaptation of this species to high nitrogen concentrationscontained in this very hazardous wastewater. In a 3 year study ofAOB community dynamics at a coke plant WWTP, the Nitrosospiralineage although present remained at the lowest relative quantityamong AOB lineages observed in this study. The Nitrosospiralineage as ‘‘K-strategist’’ grows slowly and is well adapted to lowsubstrate and oxygen conditions (Schramm et al., 1999). Thus, Nitr-osospira is often widespread in soil and terrestrial environments(Horz et al., 2004) but is less abundant in wastewater treatmentsystems (Wells et al., 2009; Kim et al., 2011a). Combinations ofcertain environmental conditions such as abundant substrate andhigh DO concentrations may repress the Nitrosospira lineage insuch a WWTP.

Throughout the study period, nitrification fluctuated while theAOB community showed resistance to change in the operationaland environmental conditions. This contrasts with oscillatorymicrobial dynamics that have been observed in other microbial

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environments. Wittebolle et al. (2005) observed that nitrificationfailures occurred as the AOB community shifts. Additionally, vari-ations in influent load affected the nitrifying community in afull-scale reactor treating tannery wastewater (Cabezas et al.,2009). Here, AOB lineages also fluctuated with operational timein the bioreactor (Wells et al., 2009). In our WWTP, the constancyof dominant AOB populations observed indicates that a climaxcommunity might be achieved. It is likely that either the strongselection pressure provided by related substrates and the ratherlong SRT contributed considerably to stabilizing the AOB commu-nity in this WWTP (Hallin et al., 2005), or the total selective pres-sure for the AOB in the full-scale WWTP was inadequate to inducea population shift (Kim et al., 2011a). Since many types of bacteriacannot withstand the continuous presence of a large amount oftoxic pollutants, it may be reasoned that some toxic componentsof the coke plant wastewater affected a lower level of AOB diver-sity. Indeed, this finding indicated that low diversity of AOB mayinduce nitrification failure. Thus, the presence of several types ofAOB of different character could prove advantageous for stabilizingthe nitrification process.

3.4. Nitrifying bacteria populations analyzed with qPCR

For all activated sludge samples collected over a 3 year periodfrom the coke plant WWTP, 16S rRNA genes for total bacteria,AOB, Nitrobacter, and Nitrospira were quantified via qPCR assays.Fig. 5 shows changes in quantities of total bacteria, AOB, Nitrobact-er, and Nitrospira in the aerobic tank of the full scale process. In allsamples, the total bacterial population in the aerobic tank rangedfrom 9.2 � 1011 to 5.9 � 1012 cells/L and remained constant withthe exception of March 2008 (1.4 � 1011 cells/L). These values arein the same order of magnitude as those obtained from activatedsludge samples from industrial WWTPs (Layton et al., 2005; Kimet al., 2011a).

The amount of AOB 16S rRNA gene averaged 7.2 � 1010 cells/L,but varied from a maximum of 3.0 � 1011 cells/L on January 2009to a minimum of 1.2 � 109 cells/L on March 2008 (Fig. 5) andwas in good agreement with the average 9.3 � 1010 cells/L ofAOB in the activated sludge from a petroleum refinery WWTP (Fig-uerola and Erijman, 2010). For the activated sludge samples whichshowed the lowest potential nitrification activity (in October 2007,March 2008, October 2008 and November 2009) (Fig. 3(f)), a rela-tive decrease in AOB population occurred compared to the sampleof the previous month. On 28 March 2008, the AOB quantity was

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ria in the full-scale activated sludge process during the study period.

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about 70-fold lower than that of the activated sludge sampled on21 January 2008 which had showed the highest potential nitrifica-tion activity. This drastically reduced quantity of AOB likely ex-plains the reduced nitrification performance, but the amount ofAOB in the rest of the samples slightly dropped in the range of onlyan approximate 2- to 10-fold decline. Moreover, certain sampleswhich demonstrated a similar drop in AOB quantity still exhibitedhigh potential nitrification activity. In previous reports, the quan-tity of AOB declined suddenly by about 2 to 3 orders of magnitude,resulting in reduced nitrification performance (Knapp and Graham,2007), while another group found that considerable numbers ofAOB cells existed even under conditions of nitrification failure (Fig-uerola and Erijman, 2010). Meanwhile, the total fractions of AOBrelative to the total bacterial population in this study varied from0.51% to 8.63% in the aerobic tank. Typically, the long SRT systems(10–20 days) have a higher percentage of AOB than the short SRTsystems, resulting from better retention of the slow growing AOBcells (Kuo et al., 2006). Despite high nitrogen loading and a longSRT of our system, only an average of 2.5% of all bacterial cells inthe activated sludge consisted of AOB. This result is consistent withthe values for activated sludge samples obtained from a systemtreating high strength hazardous wastewater (1.07–3.29%), sug-gesting that high concentrations of toxic pollutants such as phenoland cyanide might suppress the growth of AOB (Kim et al., 2011a).Contrary to expectation, the activated sludge with a high percent-age of AOB did not show higher potential nitrification activity thanthat which had a low percentage of AOB. Based on these results,the decrease in AOB cells or the AOB/bacterial ratio could notclearly explain low nitrification performance.

For NOB, Nitrobacter concentrations averaged 1.18 � 1010 cells/L and represented an average of 73% of the total NOB, while Nitro-spira averaged 2.53 � 109 cells/L and represented only 27% of theNOB community in the aerobic tank (Fig. 5). Among NOB, severalstudies demonstrated that Nitrospira is a more prevalent nitriteoxidizer in activated sludge bioreactors (Harms et al., 2003; Laytonet al., 2005). In this nitrifying system, however, Nitrobacter wasdominant over Nitrospira, explained by Nitrobacter being an r-strat-egist which thrives under high nitrite concentrations (Blackburneet al., 2007). This finding was in agreement with previous studieswhich revealed that Nitrobacter was the dominant NOB in the bio-reactors treating high strength wastewater (Figuerola and Erijman,2010; Kim et al., 2011a). Interestingly, the changing pattern in con-centrations of Nitrobacter species was similar to that of AOB, indi-cating that the Nitrobacter species might be linked to AOB. Thepercentages of the summed NOB populations within total bacteriaaveraged 0.65% and ranged from 0.07% to 2.29%; the total fractionsof Nitrobacter and Nitrospira within total bacteria were approxi-mately 0.5% and 0.14%, respectively. The relatively large amountof NOB 16S rRNA genes in this system as compared to other re-search (Kim et al., 2011a; Layton et al., 2005) suggests that NOBplayed an important role in nitrogen removal. However, no signif-icant correlation was found between the quantity of NOB and thefull nitrification state (full conversion to NO3

�).Consequently, our data showed a lack of correlation between

the quantity of nitrifying bacteria and nitrification activity. Cellquantities could be monitored using the rRNA gene but thesequantities do not correlate well with nitrification activity as a re-sult of the low growth rate and yield of nitrifying bacteria (Haw-kins et al., 2008). It is known that the molecular techniquemethod based on the rRNA gene does not always correlate withtargeted bacteria activity, as AOB have a mechanism to maintaintheir ribosome content even during relatively low activity (Wagneret al., 1995). Given that specific operational strategies have beenapplied to the full-scale WWTP, the physiological state of the nitri-fying bacterial populations might be altered without a correspond-ing change in community structure or quantity (Hallin et al., 2005).

Therefore, to correctly assess the health of a WWTP’s nitrifyingbacteria, one must simultaneously monitor both the number oftheir cells and the expression of messenger RNA that can be usedto identify shifts in their activity.

3.5. Interaction of environmental factors on the nitrifying bacteriacommunity and N-removal

To explore links between the nitrifying bacteria community aswell as nitrification performance and measured environmentaldata, PCA with extracted principal components (PCs) was con-ducted (Fig. 6). A total of 20 variables including physical, chemicaland operational parameters as well as the quantities of nitrifyingbacteria were applied to a correlation matrix which was then usedto calculate eigenvalues for each PC. The first and second PCs out ofthe 20 PCs were used to interpret the original dataset. Approxi-mately 42% of total variance in the original dataset was explainedby PC1 and PC2. Fig. 5 illustrates PC 1 and 2 loadings with temporalvariation scores of PCs; the bottom-left axes are PC loading andtop-right axes are PC scores. Eleven input variables out of the 20variables fell into three different groups (A, B and C) in which allthe loading values of the variables are greater 0.5 (i.e., positiverelationship) or less than �0.5 (i.e., negative relationship). GroupA and C are associated with PC1, while group B is related to PC2.The samples, indicated by dots, were categorized according tothe sampling date at the full-scale WWTP (Fig. 6).

For PC1, grouping influent COD, influent phenol, effluent TN andeffluent ammonia concentrations in the loadings plot suggest a sig-nificant mutually positive correlation (group A). This indicates thatphenol and COD concentrations in the coke plant wastewater couldinfluence effluent TN and ammonia concentrations of the full-scaleprocess. One can therefore speculate there is a strong link betweenphenol concentration and the inhibition of nitrification. The inhib-itory effect of phenol on nitrification is well known. Earlier studiesreported an inhibitory effect of greater than 5.5 mg/L phenol onnitrifying bacteria (Neufeld et al., 1986). Also, a full-scale activatedsludge plant showed that periods of nitrification failure coincidedwith increased phenol concentrations in petroleum refinery waste-water (Figuerola and Erijman, 2010). Particularly, it should benoted that toxicity of phenol increased over time exposure throughinhibition of nutrient transport by its binding to cell membranes(Woznica et al., 2010). This implies that prolonged exposure tophenol in the system could be lead to a severe loss of nitrifyingbacteria activity. In addition to direct phenol toxicity on nitrifica-tion, fluctuations in COD are generally acknowledged to affectnitrification in actual wastewater treatment plants. Therefore,one may reasonably assume that higher levels of biodegradable or-ganic COD directly promote growth of heterotrophic bacteria. Sincethe oxygen affinity and growth rate of nitrifying bacteria is low incomparison with heterotrophic bacteria, any subsequent increasein heterotrophic bacteria may indirectly influence nitrifying bacte-ria by competing for space, oxygen and other nutrients, resulting inrepression of nitrification (Hanaki et al., 1990). Additionally, PC1explains negative correlations between effluent nitrate and cya-nide concentrations (group C) with group A. This indicates that fullnitrification performance (conversion to NO3

�) was directly/indi-rectly influenced by influent phenol and COD concentrations. Inter-estingly, effluent cyanide concentration showed an inverserelationship with nitrification performance in this system. Thiswas caused by increased exposure to chemicals such as cohesiveagents in the post-chemical precipitation process during a periodof low nitrification performance.

For PC2, grouping quantities of total bacteria, AOB and Nitrob-acter, influent SCN� and effluent nitrite concentrations in the load-ings plot suggests a significant mutually positive correlation(group B) (Fig. 6). It is reasonable to conclude that AOB quantifying

Fig. 6. PCA biplots of operational characteristics and water quality and nitrifying bacteria populations.

188 Y.M. Kim et al. / Bioresource Technology 134 (2013) 180–189

dynamics were linked to influent SCN� concentration in the PCAanalysis. Indeed, a previous study observed that the amount ofAOB was increased approximately 3-fold as SCN� concentration in-creased 4 times (Kim et al., 2011b). Additionally, the close associa-tion between Nitrobacter and nitrite concentrations signified thestrength of their mutual correlation. This result was supportedby Huang et al. (2010) who found a positive trend between nitriteconcentrations and Nitrobacter in their PCA data. Interestingly,Nitrobacter responded positively to increase in nitrite concentra-tions, while Nitrospira showed no significant relationship to nitriteconcentrations, contrary to our expectation. Perhaps, more SCN�

inflow into the aerobic tank allowed both the amount of AOBand the nitrification rate to increase through the additional ammo-nia supplied by the biodegradation of SCN� in the aerobic tank.Subsequently, the additional nitrite generated by the AOB perfor-mance proved favorable for Nitrobacter growth which could haveprovided a selective advantage for this group over Nitrospira.

The samples marked by the numbered dots were categorizedaccording to the sampling date and demonstrated changes in thefull-scale WWTP performance over time (January 2007–November2009), even though the data showed some mixed distribution. Par-ticularly, 11 out of 18 data points had negative scores on PC1 andpositive scores on PC2, demonstrating that the concentrations ofthe variables in group B and C were relatively high during the pro-cess operation. Also, it can be seen that samples shifted from theleft (group C) to the right (group A) sides, indicating that effluentnitrate and cyanide concentrations have decreased and influentCOD and phenol and effluent TN and ammonia concentrations haveincreased. Fig. 6 shows a substantial shift toward point 8 (March2008) which is characterized by relatively low concentrations ofgroup B’s variables. This corresponds with a drastic decrease ininfluent SCN� concentration (Fig. 2(g)) and quantities of AOB andNitrobacter (Fig. 5) at point 8 (on 28 March 2008). Over time, theshift of samples to group A indicated that nitrification of the full-scale WWTP was repressed by the influence of influent COD and

phenol concentrations. Consequently the serial shifts of data pointsfor the 3 years demonstrated associations between fluctuations inoperational performance and environmental factors as well as thenitrifying bacteria communities.

4. Conclusions

Based on PCA analysis, quantities of AOB and Nitrobacter corre-lated positively with their additional substrate generated from thebiodegradation of influent SCN� in the aerobic tank. Additionally,influent COD and phenol concentrations seemed the primary rea-sons for deteriorated nitrification. The inhibitory effects of toxiccompounds by an abnormal influx of those on nitrification can beeffectively reduced by pre-treatment methods such as addition ofactivated carbon.

Acknowledgements

This work was supported by a National Research Foundation ofKorea Grant funded by the Korean Government Ministry of Educa-tion, Science and Technology (NRF-2011-357-D00140). Also, thisresearch was supported by a WCU (World Class University) pro-gram through the National Research Foundation of Korea fundedby the Ministry of Education, Science and Technology (R31-30005) and the Advanced Biomass R&D Center (ABC) of KoreaGrant funded by the Ministry of Education, Science and Technology(ABC-2010-0029800).

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