Simultaneous NH4+-N and Mn2+ removal from drinking water using a biological aerated filter system: Effects of different aeration rates

Accepted Manuscript Simultaneous NH4 +-N and Mn2+ removal from drinking water using a biological aerated filter system: Effects of different aeration rates Hassimi Abu Hasan, Siti Rozaimah Sheikh Abdullah, Siti Kartom Kamarudin, Noorhisham Tan Kofli, Nurina Anuar PII: DOI: Reference:

S1383-5866(13)00463-2 http://dx.doi.org/10.1016/j.seppur.2013.07.040 SEPPUR 11333

To appear in:

Separation and Purification Technology

Received Date: Revised Date: Accepted Date:

21 December 2012 13 June 2013 27 July 2013

Please cite this article as: H.A. Hasan, S.R.S. Abdullah, S.K. Kamarudin, N.T. Kofli, N. Anuar, Simultaneous NH4 +-N and Mn2+ removal from drinking water using a biological aerated filter system: Effects of different aeration rates, Separation and Purification Technology (2013), doi: http://dx.doi.org/10.1016/j.seppur.2013.07.040

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1

Simultaneous NH4+-N and Mn2+ removal from drinking water using a biological aerated

2

filter system: Effects of different aeration rates

3 4

Hassimi Abu Hasan*, Siti Rozaimah Sheikh Abdullah, Siti Kartom Kamarudin,

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Noorhisham Tan Kofli and Nurina Anuar

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Department of Chemical and Process Engineering, Faculty of Engineering and Built

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Environment, Universiti Kebangsaan Malaysia, 43600 UKM Bangi, Selangor, Malaysia

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*Corresponding Author: Phone: +6-03-89216407; Fax: +6-03-89216148

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Email addresses: [email protected] (Hassimi Abu Hasan), [email protected]

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(Siti Rozaimah Sheikh Abdullah)

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Abstract

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The performance of a biological aerated filter (BAF) system was studied to assess the effects

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of different aeration rates on simultaneous ammonium (NH4+-N) and manganese (Mn2+)

28

removal from drinking water. Samples of drinking water with simulated high and low

29

strengths of pollution with chemical oxygen demand (COD), NH4+-N and Mn2+ were used to

30

evaluate the bio-filtration system. For high-strength polluted drinking water, the BAF system

31

showed insignificant COD removal with increased aeration rate (AR). An AR of 2.0 L/min

32

(dissolved oxygen: 5.26 mg/L) led to higher (99.3%) removal of NH4+-N and an effluent

33

concentration below the regulated concentration limit (<1.5 mg/L). However, higher

34

manganese removal (99.1%) was achieved at an AR of 0.3 L/min (dissolved oxygen: 2.94

35

mg/L). Furthermore, for low-strength polluted drinking water, up to 98.4% of NH4+-N and

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82.9% of Mn2+ were removed simultaneously at an AR of 0.1 L/min (DO: 4.68 mg/L). The

37

best conditions for simultaneous NH4+-N and Mn2+ removal from high-strength polluted

38

drinking water were achieved at ARs of 2.0 L/min and 0.3 L/min, respectively, while their

39

removal from low-strength polluted drinking water was optimised with an AR of 0.1 L/min.

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Keywords:

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Aeration; biological aerated filter; dissolved oxygen; simultaneous ammonia and manganese

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removal; optimization of treatment

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1.

Introduction

51

In Malaysian rivers generally and in drinking water treatment plants (DWTPs) especially,

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ammonia in water, in the form of ammonium (NH4+-N), is a major problem. Levels of this

53

contaminant of over 1.5 mg/L in raw and drinking water (Table 1) exceed the Malaysian

54

regulation limits and have had deleterious impacts on DWTPs due to the creation of

55

chloramines during the chlorination process [2]. Chloramines are considered to be

56

carcinogenic and negatively impact both public health and the environment [3]. NH4+-N

57

contamination has also created problems of taste and odour in water [4], has become toxic to

58

fish, and has led to oxygen depletion and eutrophication of surface waters [5]. Although

59

manganese (Mn2+) poses only a minor problem in DWTPs and in water distribution, its

60

presence in the water distribution system over long periods of time leads to pipe clogging

61

through the oxidation of Mn2+, which precipitates Mn4+ as a black colour in the water [6]. The

62

regulated standard limits for Mn2+ in raw and drinking water are 0.2 and 0.1 mg/L (Table 1),

63

respectively. Moreover, limits for COD, NO2- and NO3- have been regulated to 10 mg/L for

64

both raw and drinking water. In Malaysia, river water is the primary source of drinking water.

65

River contamination by NH4+-N and Mn2+ often enters Malaysian drinking water treatment

66

plants. This problem has caused the DWTPs to be periodically shut down due to the inability

67

of conventional systems to treat NH4+-N and Mn2+. These shutdowns consequently lead to

68

water shortages to the human population and industries. Therefore, good quality drinking

69

water treatment systems should produce safe and good drinking water [7].

70

Currently, no specific treatments are available in Malaysian DWTPs for ammonia

71

removal [8] from drinking water. Generally, NH4+-N and Mn2+ may be removed physico-

72

chemically (by chlorination, ion exchange, or air stripping) or biologically (via an activated

73

sludge system, trickling filter or rotating biological reactor). However, physico-chemical

74

treatment is not economical due to the required addition of extra chemicals as well as high

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maintenance and operating costs. Therefore, biological removal is preferable due to its low

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operating cost and also because the volume of sludge generated is appreciably smaller and

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subsequently easier to handle [6]. Given these advantages of biological removal, this research

78

aims to design and develop a biological aerated filter (BAF) system, forming part of a

79

biological treatment process, for the simultaneous removal of NH4+-N and Mn2+.

80

The BAF system was developed in the late 1970s and the 1980s to treat wastewater

81

from the slaughterhouse and pulp-mill industries [9]. While the system is well known as one

82

of the biological methods available for the treatment of contaminants from various types of

83

wastewater [10-12], it has not been widely used for the simultaneous removal of NH4+-N and

84

Mn2+ from drinking water. Among the outstanding advantages of the BAF system are its

85

flexible reactor, its ability to perform solids separation and aerobic biological treatment and

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its small space requirement [11]. It features a small footprint with a large surface area, easy

87

construction and the ability to treat water with high organic loads [13]. Additionally, the BAF

88

system offers advantages over the standard activated sludge process, because good-quality

89

effluent that is virtually free of suspended solids is achieved wherever a slow-growing

90

biomass is involved, but a shortage of space requires high specific volumetric degradation

91

rates [10]. In conventional DWTP, the treatment system includes aeration, coagulation and

92

flocculation, sedimentation, filtration and chlorination processes. In this study, the BAF

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system is proposed as an additional system before the chlorination process because this

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treatment technology involves effective microbes to simultaneous remove NH4+-N and Mn2+.

95

Thus, by adding before the chlorination process, it is believed that the effective microbes

96

could be disinfected before water distribution system.

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It is generally known that a higher dissolved oxygen (DO) concentration enhances the

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biological nitrification rate while inhibiting the denitrification process. Conversely, at lower

99

DO concentrations the nitrification process is inhibited and the denitrification process is

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enhanced [14,15]. Furthermore, at higher DO concentrations or aeration rates, scouring of

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biofilm and reduction in removal of solids may occur [16]. Some researchers have reported

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that biological removal of manganese requires fully aerobic conditions with a DO

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concentration of over 5 mg/L to precipitate manganese [17,18].

104

It is challenging to simultaneously remove NH4+-N and Mn2+ from drinking water

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using a single treatment system because of the different oxidation-reduction potential (ORP),

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DO concentration, and pH required for oxidation of NH4+-N than that of Mn2+. The NH4+-N

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may interfere with the operation of Mn2+ removal filters because too much oxygen is

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consumed by nitrification, which results in mouldy, earthy-tasting water [19]. When drinking

109

water contains both NH4+-N and Mn2+, biological Mn2+ can only be removed after complete

110

nitrification due to the necessary evolution of the redox potential [16,20,21]. The complicated

111

removal of both pollutants in a single treatment is time-consuming and expensive to operate

112

and maintain. Oxidation and reduction processes are related to aerobic and anaerobic

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conditions that are influenced by the aeration rate (AR) supplied. When the AR supplied

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provides sufficient DO for simultaneous NH4+-N and Mn2+ removal, the ORP value increases,

115

showing the oxidation of NH4+-N and Mn2+.

116

An aeration supplement to the BAF system should supply an optimum DO

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concentration for efficient simultaneous removal of NH4+-N and Mn2+, which requires a high

118

AR. However, maintenance of a higher aeration rate consumes more energy and increases

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operating costs. Optimising aeration efficiency would make the overall process not only

120

significantly more efficient but also more economical; this is an important consideration

121

because aeration consumes most of the required power [22]. Because simultaneous removal

122

of NH4+-N and Mn2+ in DWTP is a novel application of the BAF system, the purpose of this

123

study was to investigate the effects of different aeration rates (ARs) on the performance of the

124

BAF system in removing these substances. The study employed four treatments: aerobic

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using high-strength polluted drinking water (HSPDW) (Treatment 1); anoxic using HSPDW

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(Treatment 2); aerobic using low-strength polluted drinking water (LSPDW) (Treatment 3);

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and anoxic using LSPDW (Treatment 4). The effects of different ARs were investigated in the

128

aerobic phases, while the anoxic phase was used to examine the effects of lack of aeration on

129

the system’s removal of the substances under study.

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2.

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2.1. Experimental set-up of the BAF system

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An upward flow configuration of a lab-scaled BAF system was designed using a transparent

133

polyvinyl chloride (PVC) column with a height of 1.5 m, a diameter of 0.16 m and an

134

effective working volume of 15 L. The design was based on data correlating removal

135

performances with BAF dimensions that had been obtained in a previous study [23]. A

136

schematic diagram of the BAF system is shown in Figure 1.

Materials and Methods

137

The design of the top of the column included a buffer zone of 20 cm to stop media

138

from being washed out when the system backwashed the air and water mixture. Sampling

139

ports (SPs) (SP1-6) were placed along the height of the column at 20 cm intervals to allow

140

biomass and water samplings. The BAF was partially supported with plastic media for

141

biofilm growth and attachment. The medium was made from polypropylene with a dimension

142

ratio of 0.625 and density of 888 kg/m3. The characteristics of the plastic medium are listed in

143

Table 2.

144

Simulated drinking water with different strengths of pollution was forced to flow

145

upward from an influent tank using a peristaltic pump (Masterflex, Illinois) at a flow rate of

146

580 mL/min, which was sufficient to fill the BAF column in about 30 min. At start up,

147

aerobic conditions in the BAF column were maintained by aerating at a flow rate of 300

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mL/min with the assistance of a PUMA XN2040 compressor (PUMA Industrial Co., Ltd,

149

Taiwan). An aquarium air diffuser was used to distribute the oxygen throughout the column

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to ensure homogenisation of air and water. Moreover, a backwash system was operated

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during some periods using a water pump (CPm 158, Italy) in order to prevent the reactor from

152

clogging and to maintain biofilm activity.

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2.2. Operating conditions

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The BAF system constructed for this study was operated at the ARs and air/water flow rate

155

ratios shown in Table 3. Operations proceeded batch-wise in four treatments: HSPDW under

156

aerobic conditions (Treatment 1), HSPDW under anoxic conditions (Treatment 2), LSPDW

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under aerobic conditions (Treatment 3) and LSPDW under anoxic conditions (Treatment 4).

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The study started with HSPDW at an initial AR of 0.3 L/min (phase I: day 0-9), which

159

was later increased to 0.6 (phase II: day 10-18), 1.0 (phase III: day 19-29) and 2.0 L/min

160

(phase IV: day 30-39). No aeration was provided to the HSPDW in the BAF system during

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phase V (day 40-46), resulting in a prolonged anoxic phase. In phase VI (day 47-56), a low

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AR of 0.1 L/min was set for treatment of LSPDW. Finally, phase VII (day 57-61) imposed

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anoxic conditions in the treatment of LSPDW as no aeration was provided to the BAF system.

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For clear observation of aeration effects, the HSPDW used in both treatments 1 and 2 had

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fixed concentrations of 1008 mg/L COD, 99 mg/L NH4+-N and 5.9 mg/L Mn2+. LSPDW was

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used in both treatments 3 and 4 (Table 4). Throughout the study, the treatments were executed

167

with a 24 h hydraulic retention time (HRT), a daily flow rate of 15 L/d, a pH maintained in

168

the range of 6-7, a hydraulic loading of 31.1 m3/m2.h.

169

The backwash was frequently operated every two weeks depending on the sampling

170

and simultaneous removal performance. Wherever clogging of samples was detected on the

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media, the backwash was operated to remove excess accumulated biomass. When the

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treatment of HSPDW and LSPDW was observed unstable and gradually decreased, the

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backwash operation was also performed. The backwash water flow rate was averagely set at

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10 L/min and adjusted as required through the backwash valve, while the air flow rate was

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maintained at 0.1 L/min. Backwash was performed according to the following procedures: 1)

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air scouring for about 5 min at the bottom of column, 2) simultaneous air and recycled

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backwash water for 5 min, and finally 3) air and water flows stopped and column contents

178

allowed to settle for 10 min before withdrawing the backwash water from the column. After

179

the backwash process, the removal performance of HSPDW increased and showed a stable

180

removal performance. Nevertheless, the frequency of the backwash was maintained once in

181

two weeks’ time during the treatment of LSPDW, despite the performances during the period

182

did not show any clogging and sudden deterioration in simultaneous NH4+-N and Mn2+

183

removal.

184

2.3. Simulation of contaminated drinking water

185

To simulate HSPDW and LSPDW, synthetic water was prepared with tap water containing

186

important energy sources such as carbon, nitrogen and phosphorus. The chemical

187

compositions used are listed in Table 4. LSPDWs were modelled on the normal levels of

188

COD, NH4+-N and Mn2+ in raw water, as listed in Table 5. In order to investigate the

189

performance of the BAF system at high contaminant levels, HSPDW was simulated at a level

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of pollution that is higher than that of the typical range of permissible levels of contaminated

191

drinking water. The simulation of HSPDW was based on previous literature that reported the

192

COD levels in raw water as higher as 2.1-2418 mg/L [24], 4-1794 mg/L [25], 200-1200 mg/L

193

[26] and 200-700 mg/L [27].

194

2.4. Bacterial source and identification

195

The degrading bacteria for the biofilm were taken from the sewage activated sludge (SAS)

196

water treatment plant in Putrajaya, Malaysia. This bacteria source was chosen for seeding to

197

the BAF due to its very rich content of bacteria for organic and inorganic degradation. The

198

bacteria were acclimatised and enriched in a 5 L batch reactor with sufficient energy sources

199

including glucose (200 mg/L), inorganic nutrients such as ammonia as nitrogen source (40

Page 8 of 40

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mg/L), phosphorus (2.5 mg/L), sulphur (10 mg/L) and trace elements (iron 0.3 mg/L,

201

manganese 0.3 mg/L, and magnesium 8 mg/L) for bacterial growth and enrichment. After the

202

acclimatisation process, the bacteria were transferred to the BAF system for experimental

203

studies. Sodium bicarbonate (100 mg/L) was added as an external source of alkalinity at the

204

initiation of each run in order to prevent a shortage of the inorganic carbon that nitrifying

205

bacteria require for their metabolism.

206

About 50 mL of mixed culture from BAF system was sampled and agitated to obtain

207

homogeneous suspensions. The mixed culture (10 mL) was transferred into nutrient broth

208

(100 mL) and then was shaked and incubated at 150 rpm and 37 ºC for 24 hours. After the

209

incubation, the culture was serially diluted from 10-1 to 10-5 in sterile saline water. About 0.1

210

mL of each dilution samples were spread on nutrient agar plates and incubated in a growth

211

chamber (GC 1050, Protech, Malaysia) at 37 ºC for 48 hours. The appeared cell colonies after

212

the incubation were repeated streaking with plastic loop on fresh agar plates to obtain pure

213

isolates. Afterward, bacterial DNA of each pure isolate was extracted from a respective

214

suspension in nutrient broth that was cultivated at 37oC for 24 h. The extraction was

215

conducted using Wizard® Genomic DNA Purification Kit (Promega, USA) protocol for

216

isolation of genomic DNA from Gram positive and negative bacteria. Universal primers 27F

217

(AGAGTTTGATCCTGGCTCAG) and 1492R (GGTTACCTTGTTACGACTT) were used to

218

amplify 16S RNA gene according to the PCR amplification protocol provided by Promega

219

Manufacture (USA). The PCR was performed using Mastercycler (Epgradient S, Eppendorf,

220

Version 3.608). PCR-amplified product was purified by Wizard® Plus SV Minipreps DNA

221

Purification System (Promega, USA). The PCR product was sent to First BASE Laboratories

222

Sdn. Bhd (Kuala Lumpur, Malaysia) for the 16S RNA sequencing. Finally, the result of 16S

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rRNA sequence of the isolate was compared with those of other microorganisms by way of

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BLAST

through

National

225

(http://www.ncbi.nlm.nih.gov).

Centre

for

Biotechnology

Information

homepage

226 227

2.5.

Analytical methods

228

Sampling was done at SP6 (Figure 1) after the end of the HRT. Samples were collected in 1

229

L polypropylene bottles and immediately analysed. A nitrate cellulose membrane filter (0.45

230

µm; Whatman, Germany) was used to filter the excessive mixed liquor suspended solids

231

(MLSS). Chemical oxygen demand (COD) was measured according to the HACH reactor

232

digestion method using a spectrophotometer (HACH DR/2010, USA). The ammonium

233

nitrogen (NH4+-N) and MLSS were measured by the Nesslerisation method at an absorbance

234

of 425 nm and by the gravimetric method, respectively, as described in the APHA standard

235

method [28]. Manganese in the form of total Mn was detected using an adsorption atomic

236

spectrometer, AAS (AAnalysat 800, Perkin Elmer, Massachusetts, USA). Since Mn2+ ions are

237

soluble in water [29], thus the measured total Mn given by AAS are assumed to represent the

238

soluble Mn2+ in the samples. The oxidised Mn2+, in the form of Mn4+ will be precipitated as

239

MnO2 [30]. Levels of nitrite (NO2--N) and nitrate (NO3--N) were determined through ionic

240

chromatography (IC: Model 882 Compact IC Plus, Switzerland). Probes for determination of

241

pH (Model PD1R1 GLI, USA), ORP (Model PD1R1 GLI, USA) and DO (Model 5400 GLI,

242

USA) were used daily to monitor the samples.

243

2.6. Statistical analysis

244

The results were analysed using a one-way analysis of variance (ANOVA) with a significant

245

difference of p < 0.05. Statistical calculations were executed with SPSS software for

246

Windows, version 16.0 (SPSS Inc. USA).

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3.

Results and discussion

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3.1. COD Removal

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Theoretically, increasing the AR should also increase DO levels in the water. In biological

250

treatment processes, the DO concentration in the bulk liquid has always been controlled to

251

remain above 2 mg/L to support pollutant removal [15]. With DO concentrations less than 2

252

mg/L, the system’s ability to remove pollutants biologically is inhibited. Figure 2 shows the

253

variation of effluent COD concentrations and the removal percentages under different ARs

254

and drinking water pollution strengths. During phases I to IV (0.3 to 2.0 L/min AR) under

255

aerobic conditions, the COD concentration in the effluents varied from 5 mg/L up to 98 mg/L,

256

with an insignificant change in the removal percentage from 96.1% to 97.9%. However, when

257

no aeration was provided at phases V (HSPDW) and VII (LSPDW), COD removal was only

258

44.2% and 45.3%, respectively, which resulted in effluent COD concentrations of 561 mg/L

259

and 65.2 mg/L, respectively. These results were not surprising because the DO concentration

260

at both phases were also lower at 0.04 mg/L and 0.02 mg/L (Table 6), which was not enough

261

for COD degradation.

262

In phase VI (0.1 L/min) with LSPDW, the COD effluent concentration was 9.4 mg/L

263

(below the Malaysian standard limit of 10 mg/L) with a removal percentage of 91.2%.

264

Throughout, statistical analyses indicated insignificant effects of the different ARs in phases I

265

to IV (p > 0.05). However, with no aeration (phase V and VII) and at 0.1 L/min AR (phase

266

VI), the results showed significant effects on COD removal.

267

3.2. NH4+-N removal

268

DO is a vital parameter in the nitrification process because oxygen is used by autotrophic

269

nitrifiers in the process of converting NH4+-N to nitrite (NO2--N) and then to nitrate (NO3--N).

270

However, autotrophic nitrifiers have a much lower growth rate and a higher affinity constant

271

for oxygen than do heterotrophic bacteria. Thus, autotrophic nitrifiers are outcompeted for

Page 11 of 40

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DO by heterotrophic bacteria and therefore forced to exist within the innermost area (i.e.

273

furthest from the surface) of the biofilm where the DO concentration is lower than in the

274

outermost (surface) area [15,31]. The lack of DO for the autotrophic nitrifiers in the innermost

275

area of the biofilm leads to the requirement of high DO concentrations in the BAF system to

276

achieve efficient NH4+-N removal.

277

Figure 3 shows levels of NH4+-N removal at various ARs in the BAF system. It can be

278

seen that in phases V (HSPDW) and VII (LSPDW), when no aeration was provided to the

279

BAF system, the removal of NH4+-N was only 45.3% and 5.3%, respectively, and the effluent

280

concentrations were 54.9 mg/L and 10.3 mg/L, respectively. The poor BAF performances at

281

those phases were due to the prevailing anoxic conditions throughout the column when DO

282

was detected at lower than 2.0 mg/L (Table 6). At phase I, when the AR was set to 0.3 L/min,

283

NH4+-N removal was 53.2% and the remaining NH4+-N effluent concentration was 45.9

284

mg/L. As the AR was increased to 0.6 L/min (phase II) and then to 1.0 L/min (phase III),

285

NH4+-N removal increased up to 80.8% and 88.6% with effluent concentrations of 18.7 mg/L

286

and 10.9 mg/L, respectively. At the maximum AR of 2.0 L/min (phase IV), NH4+-N was

287

efficiently removed so that the effluent concentration of 0.65 mg/L met the regulated limit.

288

The final DO concentration at this phase was 5.26 mg/L, which was much higher than those

289

found for the other phases. During phase VI, with LSPDW and a 0.1 L/min AR, the BAF

290

system performed well and resulted in 98.4% removal of NH4+-N. Therefore, it was found

291

that at either strength of NH4+-N pollution of drinking water, the BAF system can perform

292

well for NH4+-N removal with sufficient DO at the optimum ARs of 2.0 L/min for HSPDW

293

and 0.1 L/min for LSPDW. Additionally, the ANOVA revealed that the AR has a significant

294

effect on NH4+-N removal for phases I-V with an F-ratio of 33.2 (p < 0.05).

295

According to Dong et al., [15], NH4+-N can be removed well when the DO

296

concentration in the bulk liquid is high. The high requirement for DO to achieve NH4+-N

Page 12 of 40

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removal in a BAF system appears to have been triggered by the low level of oxygen in the

298

biofilm attached to the plastic media, a condition that was not suitable for nitrifier growth. As

299

reported by Dong et al., [15], more than 96% of NH4+-N was removed at a DO concentration

300

of 0.5 mg/L using a membrane aeration/filtration combined bioreactor (CMBR), but the

301

removal decreased to 69.3 % when the DO concentration decreased to 0.1 mg/L. Another

302

study reported that at a high DO concentration, almost 9 mg/L, up to 99% of NH4+-N, was

303

removed [31].

304

Figure 4 presents the nitrogen composition of the effluent across phases. It was found

305

that in phase IV using HSPDW, more NOx--N (total NO2--N and NO3--N) was formed by

306

NH4+-N conversion, with the average NOx--N concentration of the effluent reaching 142

307

mg/L. The average initial concentration of NOx--N in the influent (raw water) for all phases

308

was in range of 3.5-16.8 mg/L. Moreover, because the treatment process was conducted

309

batch-wise, the accumulation of NOx--N in the BAF system during phase IV also resulted in

310

higher NOx--N concentrations in the effluent of more than the 100 mg/L fed. On the other

311

hand, the amount of NOx--N formed in phase II was lower than those in other phases using

312

HSPDW, with average effluent NOx--N of 15.9 mg/L. Meanwhile, NOx--N formed in phases

313

VI and VII using LSPDW was less than 10 mg/L. In phases I, II and V, NOx--N was observed

314

to be greater than 10 mg/L, with concentrations of 16.7 mg/L (phase I), 15.9 mg/L (phase II),

315

and 16.6 mg/L (phase V). Because the influent contained an initial NOx--N concentration of

316

5.1 mg/L (phase I), 4.8 mg/L (phase II) and 8.5 mg/L (phase III), the effluent NOx--N

317

concentrations at the three phases were higher than the Malaysian standard limit. However,

318

when NOx--N was broken down into concentrations of NO2--N and of NO3--N, both were less

319

than the standard limit.

320

These data indicate that simultaneous nitrification and denitrification still occurred in

321

the BAF system, even at high DO concentrations (Table 6). Because the main focus of this

Page 13 of 40

322

study was to investigate the effects of AR on simultaneous NH4+-N and Mn2+ removal in a

323

BAF system used as an additional treatment in DWTP, the denitrification process that is

324

responsible for nitrate removal was not investigated.

325

3.3. Mn2+ removal

326

Biological Mn2+ removal has been widely reported to require as long as 3-4 months for the

327

establishment of a degrading biofilm on a biofilter for stable performance [16,20,21].

328

However, some have also reported that the removal could be performed within a shorter

329

period after inoculation of the bacteria [32]. According to previous studies [16,20,21],

330

biological Mn2+ oxidation can only take place after complete nitrification when water contains

331

NH4+-N because of the necessary evolution of the redox potential. In contrast, this study

332

found that biological Mn2+ oxidation occurred simultaneously with the NH4+-N removal

333

process when the DO requirement of the bacteria was met. The possibility of this

334

simultaneous removal has also been suggested in other research [33-37].

335

Figure 5 shows the result of Mn2+ removal in drinking water treatment. With an

336

influent concentration of about 5.9 mg/L, better Mn2+ removal (99.1%) was achieved at an

337

AR of 0.3 L/min (phase I) and the average effluent concentration of Mn2+ was below the

338

Malaysian standard limit (Table 6). However, increasing the AR to 0.6, 1.0 and 2.0 L/min

339

resulted in poorer effluent quality with Mn2+ concentrations of 1.72 mg/L, 1.59 mg/L and 1.60

340

mg/L, respectively. Nevertheless, while the oxidation of Mn2+ to Mn4+ still occurred in the

341

BAF system, it was not completely accomplished. The increase of Mn2+ in the effluent of

342

phases II, III and VI may also be due to the re-release of Mn4+ to Mn2+. The DO levels at

343

phase II-VI were observed to be 3.67 mg/L (phase II), 4.46 mg/L (phase III) and 5.26 mg/L

344

(phase VI), which were almost and over than 4 mg/L. Raveendran et al., [38] found that Mn4+

345

will be converted to Mn2+ when DO in the biological treatment system is over than 4 mg/L.

Page 14 of 40

346

Furthermore, the competition for DO uptake between ammonia-oxidising bacteria and

347

manganese-oxidising bacteria may also account for the poor Mn2+ removal performance as

348

AR increased [39]. At lower Mn2+ concentrations and ARs (phase VI), 96.2% removal of

349

Mn2+ was achieved with an effluent concentration of 0.08 mg/L (below the Malaysian

350

standard limit). However, under anoxic conditions (phases V and VII), Mn2+ removal was less

351

successful. The anoxic conditions were unsuitable for biological Mn2+ oxidation because the

352

degrading biofilms were not able to obtain sufficient oxygen for bacterial cells to oxidise

353

Mn2+ to Mn4+ molecules in the most energy-efficient way. Hence, it can be observed that the

354

simultaneous biological removal of NH4+-N and Mn2+ could only occur in the one-stage

355

treatment when sufficient DO is present in the BAF system. By contrast, previous studies

356

[16,20,21] reported that biological Mn2+ removal only took place after completion of the

357

nitrification process to allow for the necessary evolution of the redox potential.

358

The SAS used to seed the biofilm in the BAF system contained a wide variety of

359

microorganisms that are useful for biological NH4+-N and Mn2+ removal. From isolation and

360

identification of the SAS, six different bacterial strains with four rods and two cocci types

361

were found. The microbial community in the SAS was identified as Bacillus cereus,

362

Lysinibacillus sphaericus, Microbacterium oxydans, Bacillus thuringiensis, Staphylococcus

363

sp. and Micrococcus luteus with overall similarity of the sequence of the isolated strains with

364

the Gen Bank database was more than 95%. It has been found from previous researchers that

365

two groups of bacteria, i.e., autotrophic nitrifiers such as Nitrosomonas europaea, N.

366

halophila and N. mobilis [2] and the heterotrophic nitrifiers Pseudonocardia ammonioxydans

367

H9T [40], Acinetobacter calcoaceticus HNR [41] and Bacillus strains [42,43] were

368

responsible for biological NH4+-N removal. On the other hand, bacteria that oxidise Mn2+ to

369

Mn4+ belong to the genera Sphaerotilus, Gallionella, Leptothrix, Crenothrix, Clonothrix,

370

Metallogenium [35] and Bacillus [44,45].

Page 15 of 40

371

The capability of the isolated bacteria in oxidised Mn2+ was confirmed through a batch

372

experiment using shake flasks, which prove that biological Mn2+ removal occurred in the

373

BAF system [46,47], consequently giving evidence that Mn2+ removal occurred through

374

biological oxidation processes with sufficient DO levels. On the other hand, the reconversion

375

of Mn4+ to Mn2+ was due to the presence of B. cereus in the microbial community of SAS,

376

which it may act as manganese-reducing bacteria (MnRB) [30]. As further evidence for

377

biological Mn2+ removal in the BAF system, the initial pH value was less than pH 7. If the

378

oxidation of Mn2+ occurs through a chemical reaction, the oxidation rate has been found to be

379

extremely slow at pH levels under 8.5 [35] where an HRT of more than 24 h is required to

380

complete the elimination of Mn2+. Hence, the low influent pH of 7 suggests that Mn2+ is

381

removed through biological oxidation, consequently lowering the pH to less than pH 6 for

382

HSPDW (Table 6) and pH 6.5 for LSPDW (Table 6) in the effluent. The ANOVA showed a

383

significant effect of different ARs on Mn2+ removal (phase 1-V) with an F-ratio of 51 (p <

384

0.05), giving evidence that Mn2+ in drinking water can be removed using a BAF system with

385

optimal DO supplementation.

386

3.4. Variations of pH and ORP

387

Katsoyiannis and Zouboulis [48] have found that Mn2+ was efficiently removed to below the

388

maximum concentration limit at applied experimental conditions of 3.8 mg/L DO

389

concentration, pH 7.2 and ORP of 340 mV. However, at the end of the treatment, the initial

390

DO of 3.8 mg/L decreased to 2 mg/L while other conditions were almost constant. Effluent

391

pH and ORP for HSPDW (phase I-V) and LSPDW (phase VI-VII) for this study are presented

392

in Figure 6. In phases I to IV, the minimum pH and ORP were recorded as pH 4.1 and 329

393

mV while maximum values were pH 6.4 and 501 mV, respectively. When the water was

394

polluted at low strength, i.e. at low levels of NH4+-N and Mn2+ (phase VI, with LSPDW),

395

effluent was recorded as having pH 6.4 and ORP 346 mV. For the anoxic condition at phases

Page 16 of 40

396

V (HSPDW) and VII (LSPDW), pH levels of 6.1 and 6.4 and ORPs of -101 mV and -259

397

mV, respectively, were observed. The negative ORP value in the anoxic phase confirms the

398

lack of oxidation necessary for simultaneous biological removal of NH4+-N and Mn2+. For

399

comparison, Burger et al., [49] found that manganese was removed to below the Canadian

400

drinking water guideline value of 0.05 mg/L at pH 6.5. However, at pH 7.5, the manganese

401

effluent concentration was higher than 0.1 mg/L. Meanwhile, the optimum pH for biological

402

NH4+-N removal has been reported as pH 7-8 [50]. The difference in pH conditions for

403

removal of NH4+-N and Mn2+ showed that simultaneous removal of these pollutants is

404

difficult to achieve using a common biological treatment system.

405

3.5. Comparison with other studies

406

As listed in Table 7, a comparison of this study’s results for biological COD, NH4+-N and

407

Mn2+ removal with those of other studies revealed that, across studies, maximum COD, NH4+-

408

N and Mn2+ removal only occurred in the presence of an optimum DO concentration in the

409

water. However, while a high DO concentration is necessary for removal of these substances,

410

it is not necessarily sufficient for good removal performance. This study showed that COD

411

and NH4+-N were completely removed to the lower concentration when the DO concentration

412

was above 5 mg/L, but Mn2+ was not as effectively removed when COD and NH4+-N were

413

removed at higher DO levels.

414

Similar phenomena in terms of DO levels were also observed by Liu et al., [51], who

415

found that at DO concentrations over 5 mg/L (air:water ratio = 6:1), NH4+-N removal was

416

inhibited due to the nitrifiers being harmed by the higher air flow rate. The essential level of

417

DO for nitrification and COD removal was also studied by Dong et al.,[15], who studied

418

drinking water treatment using a membrane aeration/filtration combined bioreactor. They

419

observed maximum COD and NH4+-N removal at a DO concentration of 0.5 mg/L with

420

94.5% and 96% removal, respectively. The low DO requirement was possible because the

Page 17 of 40

421

biofilm attachment on the membrane contained a high level of oxygen, which is favourable

422

for nitrifier growth. Although nitrifiers consumed low levels of DO in the membrane

423

aeration/filtration combined bioreactor, the cost of the system was still high because of the

424

high cost and maintenance requirements of the membrane [52].

425

Despite extensive research on COD and NH4+-N removal, little investigation has been

426

done on simultaneous biological NH4+-N and Mn2+ removal in terms of DO influence or ARs.

427

Tekerlekopoulou and Vayenas [53] investigated the simultaneous removal of biological

428

ammonia, iron and manganese from potable water using a trickling filter. However, the study

429

focused on the effects of different feed concentrations on the interactions among removal

430

levels of ammonia, iron and manganese, which resulted in various ORP values along the filter

431

depth. Katsoyiannis and Zoubouli [48] studied biological manganese and iron removal from

432

groundwater by a fixed bed filter, where manganese present in the groundwater was reduced

433

to below the maximum concentration limit given an initial DO concentration of 0.9 mg/L.

434

During their treatment, the DO first increased to 3.8 mg/L at the aeration stage and later

435

decreased to 2.0 mg/L in the effluent. This decrease occurred because the DO was consumed

436

by manganese-oxidising bacteria to oxidise the soluble Mn2+, thus precipitating Mn4+ as black

437

colouration in the water.

438

Table 8 summarises the findings of our study for the optimum AR for simultaneous

439

NH4+-N and Mn2+ removal from drinking water using a BAF system. Simultaneous removal

440

of these substances occurred efficiently for LSPDW at an optimum AR of 0.1 L/min with an

441

effluent DO level of 4.68 mg/L. When a lower AR provides sufficient oxygen to the BAF

442

system, consumption of electrical energy can be reduced and the effectiveness of drinking

443

water treatment plants enhanced.

444

Page 18 of 40

445

4.

Conclusions

446

A lab-scale BAF system was developed and tested to investigate the effect of ARs on

447

simultaneous removal of NH4+-N and Mn2+ from drinking water. The main conclusions from

448

this study are as follows: increasing the AR (Treatment 1: HSPDW) yields no significant

449

effect on COD removal; however, significant removal of NH4+-N was observed as the AR

450

increased, resulting in effluent concentrations of NH4+-N below the regulated limit. Mn2+

451

removal decreased when the AR increased from 0.6 to 2.0 L/min. The results of Treatments 2

452

(HSPDW) and 4 (LSPDW) lead to the conclusion that for the treatment of both heavily and

453

lightly polluted drinking water, DO levels in the BAF system should be maintained at 2-4

454

mg/L for effective removal of NH4+-N and Mn2+. For treatment of lightly polluted drinking

455

water (LSPDW; Treatment 3), even at the slow AR of 0.1 L/min, NH4+-N and Mn2+ were

456

removed simultaneously at high removal percentages. At this stage, the DO concentration in

457

the effluent was observed to be over 4 mg/L. The optimum ARs for maximum removal of

458

NH4+-N and Mn2+ from HSPDW were 2.0 L/min and 0.3 L/min, respectively. However, an

459

AR of only 0.1 L/min was required for treatment of LSPDW. Finally, with sufficient DO

460

levels, the BAF system performed well for removal of the studied substances from both

461

HSPDW and LSPDW. The use of the minimum AR, which is sufficient for simultaneous

462

NH4+-N and Mn2+ removal from lightly polluted drinking water, will result in lower electricity

463

consumption and operation costs as less energy is required for aeration.

464 465

Acknowledgement

466

This research was financially supported by the Ministry of Science, Technology and

467

Innovation, Malaysia (MOSTI) through grant number 02-01-02-SF0367.

468 469

Page 19 of 40

470

References

471

[1]

NSDWQ, National Standard for Drinking Water Quality, Engineering Service Division,

472

Malaysian

473

http://kmam.moh.gov.my/public-user/drinking-water-quality-standard.html.

474

[2]

Ministry

of

Health,

Putrajaya,

Malaysia,

2004.

E. Okoniewska, J. Lach, M. Kacprzak, E. Neczaj, The removal of manganese, iron and

475

ammonium nitrogen on impregnated activated carbon, Desalination 206 (2007) 251-

476

258.

477

[3]

478 479

McGraw-Hill, New York, USA, 2004. [4]

480 481

[5]

D.V. Vayenas, S. Pavlou, G. Lyberatos, Development of a dynamic model describing nitrification and nitratification in trickling filters. Water Res. 31 (1997) 1135-1147.

[6]

484 485

W.R. Markesbery, W.D. Ehmann, M. Alaudin, T.I.M. Hossain, Brain trace element concentrations in aging, Neurobiol. Aging 5 (1984) 19-28.

482 483

Metcalf and Eddy, 2004. Wastewater Engineering, Treatment and Reuse (4th ed.).

A.G. Tekerlekopoulou, D.V. Vayenas, Ammonia, iron and manganese removal from potable water using trickling filters, Desalination 210 (2007) 225-235.

[7]

H.A. Hasan, S.R.S. Abdullah, S.K. Kamarudin, N.T. Kofli, Ammonia and manganese

486

problems in Malaysian drinking water treatment, World Appl. Sci. J. 12 (2011) 1890-

487

1896.

488

[8] SPAN, Activities report by National Water Services Commission: Ammonia treatment

489

plant will be proposed to the cabinet. [Laporan aktiviti Suruhanjaya Perkhidmatan Air

490

Negara: cadangan loji rawatan ammonia akan dibawa ke kabinet]. No. 6/09, pg. 3, 2009.

491

http://www.span.gov.my/pdf/events/activityAugust09.pdf

492 493

[9]

W.S. Chang, S.W. Hong, J. Park, Effect of zeolite media for the treatment of textile wastewater in a biological aerated filter, Process Biochem. 37 (2002) 693-698.

Page 20 of 40

494

[10] E. Rother, P. Cornel, A. Ante, P. Kleinert, R. Brambach, Comparison of combined and

495

separated

biological

aerated

filter

(BAF)

performance

for

pre-

496

denitrification/nitrification of municipal wastewater, Water Sci. Technol. 46 (2002)

497

149-158.

498

[11] D. Su, J. Wang, K. Liu, D. Zhou, Kinetic performance of oil-field produced water

499

treatment by biological aerated filter, Chinese J. Chem. Eng. 15(4) (2007) 591-594.

500

[12] F. Liu, C.C. Zhao, D.F. Zhao, L.H. Liu, Tertiary treatment of textile wastewater with

501

combined media biological aerated filter (CMBAF) at different hydraulic loadings and

502

dissolved oxygen concentrations. J. Hazard. Mater. 160 (2008) 161-167.

503 504 505 506

[13] A.T. Mann, T. Stephenson, Modelling of biological aerated filters for wastewater treatment, Water Res. 31(10) (1997) 2443-2448. [14] K. Pochana, J. Keller, Study of factors affecting simultaneous nitrification and denitrification (SND), Water Sci. Technol. 39 (1999) 61-68.

507

[15] W.Y. Dong, H.J. Wang, W.G. Li, W.C. Ying, G.H. Gan, Y. Yang, Effect of DO on

508

simultaneous removal of carbon and nitrogen by a membrane aeration/filtration

509

combined bioreactor, J. Membrane Sci. 344 (2009) 219-224.

510 511 512 513

[16] S.L. Harris, T. Stephenson, P. Pearce, Aeration investigation of biological aerated filters using off-gas analysis, Water Sci. Technol. 34(3-4) (1996) 307-314. [17] P. Mouchet, From conventional to biological removal of iron and manganese in France, J. Am. Water Works Assoc. 84 (1992) 158-166.

514

[18] V.A. Pacini, A.M. Ingallinella, G. Sanguinetti, Removal of iron and manganese using

515

biological roughing up flow filtration technology, Water Res. 39 (2005) 4463-4475.

516

[19] WHO, Ammonia in drinking water: In Guidelines for drinking-water quality (second

517

ed., volume 2), Health criteria and other supporting information. World Health

518

Organization, Geneva, 1996.

Page 21 of 40

519

[20] H. Frischhertz, F. Zibuschka, H. Jung, W. Zerobin, Biological elimination of iron and

520

manganese, Water Supply, 3 (1985) 125-136.

521

[21] J. Vandenabeele, M.V. Woestyne, F. Houwen, R. Germonrpe, D. Vandesande, W.

522

Verstraete, Role of autotrophic nitrifiers in biological manganese removal from

523

groundwater containing manganese and ammonium, Microbiol. Ecol. 29 (1995) 83-98.

524

[22] L. Mendoza-Espinosa, T. Stephenson, A review of biological aerated filters (BAFs) for

525

wastewater treatment, Environ. Eng. Sc. 16(3) (1999) 201-216.

526

[23] H.A. Hasan, S.R.S. Abdullah, S.K. Kamarudin, N.T. Kofli, A review on the design

527

criteria of biological aerated filter for COD, ammonia and manganese removal in

528

drinking water treatment, J. Institut. Eng. Malaysia 70 (2009) 25-33.

529

[24]

V. Thanapalasingam, Pollution status of the Sungai Skudai river system through heavy

530

metals. Thesis Master of Science, Faculty of Science, Universiti Teknologi Malaysia,

531

pp. 1-280, 2005.

532

[25] R.M. Bhardwaj, Water quality monitoring in India-Achievements and constraints.

533

International Work Session on Water Statistics, Vienna, pp. 1-12, 2005.

534

[26] F.N.M. Saad, N.N.N.R. Rahman, M.O.A. Kadir, F.M. Omar, Identification of pollution

535

sources within the Sungai Pinang river basin. in: H. Mokhlis (Ed.) Proceeding of the

536

Malaysian Research Group International Conference, Manchester, United Kingdom, pp.

537

478-485, 2006.

538

[27]

O. Richard, Deteriorating Water Quality in the Lake Victoria Inner Murchison Bay

539

and its impact on the drinking water supply for Kampala, Uganda, 2011. Online at:

540

http://wldb.ilec.or.jp/data/ilec/WLC13_Papers/others/3.pdf.

541 542

[28] APHA, Standard Methods for the Examination of Water and Wastewater (21st ed.), American Public Health Association, Washington, DC, USA, 2005.

Page 22 of 40

543

[29] C.K. Hope, T.R. Bott, Laboratory modelling of manganese biofiltration using biofilms

544 545

of Leptothrix discophora. Water Res. 38 (2004) 1853-1861 [30]

J.M. Cerrato, J.O.3rd. Falkinham, A.M. Dietrich, W.R. Knocke, C.W. McKinney, A.

546

Pruden, Manganese-oxidizing and -reducing microorganisms isolated from biofilms in

547

chlorinated drinking water systems, Water Res. 44 (2010) 3935-3945.

548

[31] X. Yu, Z. Qi, Z. Zhang, P. Yu, B. Liu, L. Zhang, L. Fud, Nitrogen loss and oxygen

549

paradox in full-scale biofiltration for drinking water treatment, Water Research 41

550

(2007) 1455-1464.

551

[32] T. Stembal, M. Markic, F. Briski, L. Sipos, Rapid start-up of biofilters for removal of

552

ammonium, iron and manganese from ground water, J. Water Supply: Res. Technol.-

553

AQUA, 53(7) (2004) 509-518.

554 555

[33] T. Peitchev, Characteristics of the processes in rapid sand filters used by the treatment of ground and terrace waters. Water Supply 9 (1991) S47-S51.

556

[34] A. Gouzinis, N. Kosmidis, D.V. Vayenas, G. Lyberatos, Removal of Mn and

557

simultaneous removal of NH3, Fe, and Mn from potable water using a trickling filter,

558

Water Res. 32 (1998) 2442-2450.

559

[35] T. Stembal, M. Markic, N. Ribicic, F. Briski, L. Sipos, Removal of ammonia, iron and

560

manganese from ground waters of northern Croatia-pilot plant studies, Process

561

Biochem. 40 (2005) 327-335.

562

[36] H.A. Hasan, S.R.S. Abdullah, S.K. Kamarudin, N.T. Kofli, Recognition of relevant

563

ORP, pH and DO bending points in ammonia removal from drinking water through

564

online BAF system. Inter. J. Chem. Eng., vol. 2010, Article ID 587032, 7 pages, (2010).

565

doi:10.1155/2010/587032.

566 567

[37] H.A. Hasan, S.R.S. Abdullah, S.K. Kamarudin, N.T. Kofli, Effect of organic carbon loading (OCL) on simultaneous NH4+-N and Mn

2+

removal in drinking water using a

Page 23 of 40

568

BAF system, Environ. Eng. Manage. J. 10(11) (2011) 1733-1742.

569

[38] R. Raveendran, B. Ashworth, B. Chatelier, Manganese removal in drinking water

570

systems. 64th Annual Water Industry Engineers and Operators’ Conference, All

571

Seasons International Hotel – Bendigo, Australia, pp. 92-100, 2001.

572

[39] D. Li, J. Zhang, H. Wang, H. Yang, B. Wang, Operational performance of biological

573

treatment plant for iron and manganese removal, J. Water Supply: Res. Technol.

574

AQUA, 54(1) (2005) 15-24.

575

[40] N. Ahmad, H.Y. Xu, L.P. Chen, Z.P. Liu, S.J. Liu, Enhanced biological nutrient

576

removal by the alliance of a heterotrophic nitrifying strain with a nitrogen removing

577

ecosystem, J. Environ. Sci. 20 (2008) 216-223.

578 579

[41] B. Zhao, Y.L. He, J. Hughes, X.F. Zhang, Heterotrophic nitrogen removal by a newly isolated Acinetobacter calcoaceticus HNR, Bioresour. Technol. 101 (2010) 5194-5200.

580

[42] J.K. Kim, K.J. Park, K.S. Cho, S.W. Nam, T.J. Park, R. Bajpai, (2005). Aerobic

581

nitrification-denitrification by heterotrophic Bacillus strains. Bioresour. Technol. 96

582

(2005) 1897-1906.

583

[43] H.A. Hasan, S.R.S. Abdullah, N.T. Kofli, S.K. Kamarudin, Removal of NH4+-N ion in

584

drinking water treatment using locally isolated heterotrophic nitrifier, Desalin. Water

585

Treat. 50 (2012) 294-301.

586

[44] J.P.M. de Vrind, E.W. de Vrind-de Jong, J.W.H. de Voogt, P. Westbroek, F.C. Boogerd,

587

R.A. Rosson, Manganese oxidation by spores and spore coats of a marine Bacillus

588

species, Applied Environ. Microbiol. 52 (1986) 1096-1100.

589

[45] H.A. Hasan, S.R.S. Abdullah, N.T. Kofli, S.K. Kamarudin, Isotherm equilibriums of

590

Mn2+ biosorption in drinking water by local isolated Bacillus species and sewage

591

activated sludge, J. Environ. Manage. 111 (2012) 34-43.

Page 24 of 40

592

[46] H.A. Hasan, S.R.S. Abdullah, N.T. Kofli, S.K. Kamarudin, Biosorption of manganese

593 594

in drinking water by isolated bacteria, J. Appl. Sci. 10(21) (2010) 2653-2657. [47]

H.A. Hasan, S.R.S. Abdullah, N.T. Kofli, S.K. Kamarudin, Effective microbes for

595

simultaneous bio-oxidation of ammonia and manganese in biological aerated filter

596

system, Bioresour. Technol. 124 (2012) 355-363.

597

[48] I.A. Katsoyiannis, A.I. Zouboulis, Biological treatment of Mn(II) and Fe(II) containing

598

groundwater: kinetic considerations and product characterization, Water Res. (2004) 38

599

1922-1932.

600

[49] M.S. Burger, S.S. Mercer, G.D. Shupe, G.A. Gagnon, Manganese removal during

601 602

bench-scale biofiltration, Water Res. 42 (2008) 4733-4742. [50] S. Villaverde, P.G. Garcia-Encina, F. Fdz-Polanco, Influence of pH over nitrifying

603

biofilm activity in submerged biofilters, Water Res. 31(5) (1997) 1180-1186.

604

[51] B. Liu, D. Yan, Q. Wang, S. Li, S. Yang, W. Wu, Feasibility of a two-stage biological

605

aerated filter for depth processing of electroplating-wastewater, Bioresour. Technol. 100

606

(2009) 3891-3896.

607

[52] C. Wisniewski, Membrane bioreactor for water reuse, Desalination 203 (2007) 15-19.

608

[53] A.G. Tekerlekopoulou, D.V. Vayenas, Simultaneous biological removal of ammonia,

609

iron and manganese from potable water using a trickling filter. Biochem. Eng. J. 39

610

(2008) 215-220.

611 612 613

.

614 615 616 Page 25 of 40

617

Figure Captions

618 619

Fig. 1 Schematic diagram of the biological aerated filter (BAF) system

620

Fig. 2 COD effluent concentrations and removal percentages

621

Fig. 3 NH4+-N effluent concentrations and removal percentages

622

Fig. 4 Nitrogen composition in the effluent across phases

623

Fig. 5 Mn2+ effluent concentrations and removal percentages

624

Fig. 6 Variation of pH and ORP under different ARs

625 626 627 628 629 630 631 632 633

Page 26 of 40

ORP probe DO probe pH probe SP7 Buffer zone SP6 NH4+ probe

SP5 SP4 SP3

Online monitoring

A/D converter

SP2

NO3probe

SP1

Flow meter Peristaltic pump

Influent tank

Effluent tank

Compressor Backwash pump

634 635

Fig. 1

636 637 638 639 640 641

Page 27 of 40

100%

1000

80%

800

VII

V I

600

II

III

60%

VI

IV

40% Influent

400

Removal (%)

COD concentration (mg/L)

1200

Effluent

200

20%

Removal

0%

0 0

5

10

15

20

25

30

35

642

Run (days)

643

Fig. 2

40

45

50

55

60

65

644 645 646 647 648 649 650 651

Page 28 of 40

100%

100

80%

80

V

VI

VII

60%

60 II

III

IV 40%

Influent

40

Removal (%)

NH4+-N concentration (mg/L)

120

Effluent 20%

Removal

20 I 0

0% 0

5

10

15

20

25

30

35

652

Run (days)

653

Fig. 3

40

45

50

55

60

65

654 655 656 657 658 659 660 661 662

Page 29 of 40

180 160

NOx--N (mg/)

140 120 100

I

II

III

V

IV

VI

VII

80 Influent

60

Effluent

40 20 0 0

5

10

15

20

25

30

35

663

Run (days)

664

Fig. 4

40

45

50

55

60

65

665 666 667 668 669 670 671 672 673 674

Page 30 of 40

100%

7 80%

Influent

6

Effluent 5

Removal

60%

4 I

II

III

IV

VI

V

3

40%

2

VII

Removal (%)

Mn2+ concentration (mg/L)

8

20%

1 0

0% 0

5

10

15

20

25

30

35

40

675

Run (days)

676

Fig. 5

45

50

55

60

65

677 678 679 680 681 682 683 684

Page 31 of 40

8

600

7

500 400

6

pH

5

200

4

100 I

II

III

IV

V

VI

VII

0

3

ORP (mV)

300

-100 2 1

pH

-200

ORP

-300 -400

0 0

5

10

15

20

25

30

35

40

685

Run (days)

686

Fig. 6

45

50

55

60

65

687 688 689 690 691 692 693 694 695 696 697 698 699 700 701 702 703 704 Page 32 of 40

705

Table 1 Malaysian standard limits for water quality [1] Contaminants

Raw water

Drinking water

(mg/L)

(mg/L)

COD

10

10

NH4+-N

1.5

1.5

NO2-

10

10

NO3-

10

10

Mn2+

0.2

0.1

706 707 708 709 710 711 712 713 714 715 716 717 718 719

Page 33 of 40

720

Table 2 Characteristics of plastic medium Characteristics

Value

Diameter (mm)

16

Height (mm)

10

Weight (g)

0.65

Density (kg/m3)

888

Specific surface area (m2/m3)

450

721 722 723 724 725 726 727 728 729 730 731 732 733 734 735

Page 34 of 40

Table 3 Experimental procedures for simultaneous NH4+-N and Mn2+ removal

736 Treatments

Phases

Loading rate (kg/m3.d)

Strength

Operating

Influent (mg/L)

types

conditions

COD

NH4+-N

Mn2+

COD

NH4+-N

Mn2+

Aeration (L/min)

Air/water flow rate ratios

1

I, II, III, V

HSPDW

Aerobic

1008 ± 45

99 ± 4.9

5.9 ± 0.3

1.01

0.01

0.006

0.3, 0.6, 1.0 and 2.0

29, 58, 96 and 192

2

V

HSPDW

Anoxic

1008 ± 45

99 ± 4.9

5.9 ± 0.3

1.01

0.01

0.006

0

0

3

VI

LSPDW

Aerobic

107 ± 2.6

9.8 ± 0.3

2.0 ± 0.04

0.11

0.001

0.002

0.1

9.6

4

VII

LSPDW

Anoxic

119 ± 1.6

9.8 ± 0.3

1.0 ± 0.03

0.12

0.001

0.001

0

0

737 738 739 740 741

Page 35 of 40

742

Table 4 Chemical composition of simulated drinking water contaminated at two levels Added chemicals

Glucose, C6H12O6

Chemical brand

Systerm

HSPDW

LSPDW

(mg/L)

(mg/L)

1008 ± 45

107 ± 2.6

99 ± 4.9

9.8 ± 0.3

100

100

8

8

0.3

0.3

5.9 ± 0.3

2.0 ± 0.04

4.5

4.5

2.5

2.5

(Malaysia) Ammonium sulphate,

Systerm

(NH4)2SO4

(Malaysia)

Sodium bicarbonate, NaHCO3

R & M Chemicals (United Kingdom)

Magnesium chloride,

R & M Chemicals

MgCl2.6H2O

(United Kingdom)

Iron chloride, FeCl3.6H2O

R & M Chemicals (United Kingdom)

Manganese chloride,

R & M Chemicals

MnCl2.4H2O

(United Kingdom)

Calcium chloride, CaCl2.2H2O

Systerm (Malaysia)

Potassium dihydrogen

Merck (Germany)

phosphate, KH2PO4 743 744 Page 36 of 40

745

Table 5 Characteristics of drinking water source Contaminants

Range of levels (mg/L)

COD

2.5-101

NH4+-N

1.8-12.7

Mn2+

0.003-0.520

746

Page 37 of 40

747

Table 6 Ammonium and manganese removal with different ARs Phases

Conditions

AR (L/min)

DO (mg/L)

pH

ORP (mV)

Effluent (mg/L) COD

NH4+-N

COD

NH4+-N

Mn2+

COD

NH4+-N

Mn2+

106.4

33.2

51

-

-

-

2.94

5.34

396

28.9

45.89

a

0.05

97.0

53.2

99.1

II (day 10-18) III (day 19-29)

0.6

3.67

4.38

473

33.4

18.67

1.72

96.7

80.8

69.4

1.0

4.46

5.37

431

41.1

10.95

1.59

96.0

88.6

73.8

IV (day 30-39)

2.0

5.26

5.80

427

31.2

0.65*

1.60

96.9

99.3

73.4

560.7

54.29

4.45

44.2

45.3

26.5

a

a

0.08

91.3

98.4

96.2

3.2

45.3

5.3

0

Aerobic (HSPDW)

V (day 40-46)

Anoxic (HSPDW)

0

0.04

5.89

-114

VI (day 47-56)

Aerobic (LSPDW)

0.1

4.68

6.40

346

a

VII (day 57-61)

Anoxic (LSPDW)

0

0.02

6.43

-259

65.2

9.4

0.15

10.3

F values (p < 0.05)**

Mn2+

0.3

I (day 1-9)

b

Average removal (%)

748

a

Below standard limit (< 1.5 mg/L: NH4+-N, < 0.1 mg/L: Mn2+)

749

b

F values calculated only for HSPDW. The values for LSPDW cannot be calculated as only two aeration rates were investigated.

750 751 752 753 754 755 756 757 Page 38 of 40

Table 7 Cross-study comparisons of maximum COD, NH4+-N and Mn2+ removal in terms of condition requirements

758 References

This study (2013) Dong et al., [15]

Katsoyiannis and Zouboulis [48] Liu et al., [12] Liu et al., [51]

Treatment systems BAF Membrane aeration/filtration combined bioreactor Fixed-bed filter

Wastewater types Synthetic polluted drinking water Drinking water

DO (mg/L) a 2.9 b 4.7 0.5

2.0

Groundwater

pH 5.34 6.40 7.5-8.0

ORP (mV) a 396 b 346 -

COD a 97.0 b 91.3 94.5

7.2

340

-

a

b

Removal (%) NH4+-N a 53.2 b 98.4 96

Mn2+ a 99.1 b 82.9 -

-

6.1

7.2

-

53

88

> 88 (0.05) -

3.0 (4:1)

7.8

-

90.1

92.5

-

c

Combined media BAF Two stages BAF

Textile wastewater Electroplating wastewater

759

a

high-strength polluted drinking water (HSPDW)

760

b

low-strength polluted drinking water (LSPDW)

761

c

below the European Commission maximum concentration limit

762

d

air/water ratio

d

Page 39 of 40

763

Table 8 Optimum ARs for simultaneous NH4+-N and Mn2+ removal Parameters

HSPDW

LSPDW

AR (L/min)

DO (mg/L)

COD

0.3

2.94

NH4+-N

2.0

5.26

Mn2+

0.3

2.94

AR (L/min)

DO (mg/L)

0.1

4.68

764

Page 40 of 40