REVIEW ARTICLE

Effective Biological Nitrogen Removal Treatment Processesfor Domestic Wastewaters with Low C=N Ratios: A Review

Sheng-Peng Sun,1,2,* Carles Pellicer i Nacher,2 Brian Merkey,2 Qi Zhou,1

Si-Qing Xia,1 Dian-Hai Yang,1 Jian-Hui Sun,3 and Barth F. Smets2

1State Key Laboratory of Pollution Control and Resources Reuse, College of Environmental Science and Engineering,

Tongji University, Shanghai, People’s Republic of China.2Department of Environmental Engineering, Technical University of Denmark, Kgs. Lyngby, Denmark.

3Henan Key Laboratory for Environmental Pollution Control, College of Chemistry and Environmental Sciences,

Henan Normal University, Xinxiang, Henan, People’s Republic of China.

Received: March 22, 2009 Accepted in revised form: October 28, 2009

Abstract

Discharge of nitrogenous components to water bodies can cause eutrophication, deterioration of water quality,toxicity to aquatic life, and pose a potential hazard to human and animal health. Biological nitrogen removal canremove nitrogenous components via conversion to harmless nitrogen gas with high efficiency and relative low costs.However, the removal of nitrogen from domestic wastewater with a low carbon=nitrogen (C=N) ratio can often belimited in municipal wastewater plants (WWTPs) because organic carbon is a limiting factor for denitrification. Thepresent work reviews innovative bacterial nitrogen removal pathways such as shortcut nitrification=denitrification,simultaneous nitrification=denitrification, and the nitritation Anammox process, which can remove nitrogen withlow or zero dosage of organic carbon sources. We conclude that advanced process control and some new biologicaltreatment processes including the modified anaerobic=anoxic=oxic (A2=O) process, the step-feed multistage an-aerobic=oxic (A=O) process, and new reactors like the membrane bioreactors (MBRs) and the membrane-aeratedbiofilm reactors (MABRs) can support the innovative biological nitrogen removal pathways. They can effectivelybe used for nitrogen removal from low C=N domestic wastewater without external carbon addition. In addition,conventional and alternative carbon sources for enhanced biological nitrogen removal were also reviewed. Weconclude that alternative carbon sources such as wine distillery effluent, the leachate of food waste, digested piggerymanure, hydrolyzed molasses, biologically hydrolyzed or mechanically disintegrated sludge offer the same or betterperformance for nitrogen removal at reduced costs. Finally, we suggest that (1) these new processes and technol-ogies are implemented at large scale for nitrogen removal from low C=N domestic wastewater, (2) further methodlogic are explored to introduce the Anammox pathway into domestic wastewater treatment, and (3) alternativecarbon sources are explored and optimized for supporting the denitrification. With these efforts, cost-effectivenitrogen removal from low C=N ratio domestic wastewater can be obtained in the near future.

Key words: nitrogen removal; C=N ratio; nitrification; denitrification; Anammox; carbon source; domesticwastewater; WWTP

Introduction

Human-induced eutrophication occurs due to thedischarge of high nutrients loads. Eutrophication re-

duces water quality, alters the ecological structure and func-tion of freshwater, and poses many potential hazards tohuman and animal health (Ahn, 2006; Dodds et al., 2009).

Environmental protection has led to the implementation ofcontinuously increasing wastewater treatment plant (WWTP)discharge standards for nutrients and has impelled efforts toachieve better effluent quality especially with regard to ni-trogen (Henze et al., 2008).

Biological nitrogen removal is preferred over physicochem-ical processes because it is capable of removing fixed nitroge-nous compounds to harmless dinitrogen gas (N2) in a moreeffective and economical way (Ahn, 2006; Kim et al., 2008).Nitrogenous matter in wastewater can be divided into twocategories: nonbiodegradable and biodegradable. The nonbio-degradable nitrogenous fraction is associated with the nonbio-degradable particulate chemicical oxygen demand (COD). The

*Corresponding author: State Key Laboratory of Pollution Con-trol and Resources Reuse, College of Environmental Science andEngineering, Tongji University, Shanghai 200092, People’s Republicof China. Phone: þ86 21 6598 2692; Fax: þ86 21 6598 2689; E-mail:sunsp@yahoo.cn

ENVIRONMENTAL ENGINEERING SCIENCEVolume 27, Number 2, 2010ª Mary Ann Liebert, Inc.DOI: 10.1089=ees.2009.0100

111

biodegradable nitrogenous matter is subdivided into ammonia(NHþ4 ), soluble organic nitrogen, and particulate organic ni-trogen. Particulate organic nitrogen is hydrolyzed to solubleorganic nitrogen. The soluble organic nitrogen is acted on byheterotrophic bacteria and converted to NHþ4 . The NHþ

4

serves as the nitrogen supply for synthesis of heterotrophicbiomass and as the energy supply for growth of autotrophicnitrifying bacteria (Henze et al., 2000). Presently, the mostcommonly used approach for nitrogen removal in WWTPscomprises a combination of aerobic autotrophic nitrificationof NHþ

4 to nitrite (NO�2 ) and nitrate (NO�3 ) and later anoxicheterotrophic denitrification of the oxidized N-species to N2

(Pynaert et al., 2004). Biological denitrification is carried outby heterotrophic bacteria (HB), which require a biodegrad-able organic carbon source as an electron donor. Thus, thedenitrification potential of wastewater is primarily a functionof the available organic carbon (usually expressed as biode-gradable chemical oxygen demand (bCOD)=nitrogen (N)[carbon=nitrogen (C=N)] or biochemical oxygen demand(BOD)=N ratio (Kujawa and Klapwijk, 1999). Without exter-nal carbon sources, the specific denitrification rates (SDNR),range from 0.2 to 0.6 mg-N=g-mixed liquor volatile sus-pended solids (MLVSS)=h due to endogenous processes(Kujawa and Klapwijk, 1999). Beccari et al. (1983) suggestedminimum C=N ratios of 13, and a similar value was inferredwith full-scale predenitrification experience (Komorowska-Kaufman et al., 2006). Henze et al. (1994) calculated that aminimum C=N ratio of 3.5–4.5 g-COD=g-N and proposedthat 6–11 g-COD=g-N could achieve good nitrogen removalin single sludge systems with predenitrification. The C=Nratio of domestic wastewaters is, however, often lower thanthese prescribed values, and nitrogen removal is limited bythe lack of available organic carbon source (Ryu and Lee,2009). It is desired for many municipal WWTPs to look forpractical and cost-effective alternative technologies for nitro-gen removal.

Given the mentioned references above and the C=N ratio oftypical domestic wastewater (10.5–12.5) (Tchobanoglous et al.,2004), we define a low C=N ratio domestic wastewater as awastewater with a C=N ratio lower than 8.0. To get satisfac-tory nitrogen removal performance for wastewater with aC=N ratio lower than the critical value, one way is to intro-duce an innovative nitrogen removal pathway or treatmentprocesses that can support nitrogen removal with a low orzero organic carbon demand; the other way is to add externalcarbon for denitrification. Therefore, this article reviews in-novative nitrogen removal pathways and available biologicaltreatment processes that are capable of nitrogen removal withlow or zero organic carbon demand. In addition, conventionaland alternative external carbon sources that can be used forenhanced nitrogen removal are also discussed and summa-rized. Hopefully, this review can provide guidance to mu-nicipal WWTPs operators and planners to select cost-effectivenitrogen removal alternatives for the treatment of low C=Nratio domestic wastewater.

Feasible N-Removal Pathway for Wastewaterwith Low C/N Ratio

During the past 30 years several innovative nitrogenremoval pathways have been discovered such as shortcutnitrification=denitrification, simultaneous nitrification=

denitrification, and anaerobic ammonium oxidation (Ana-mmox) (Ahn, 2006; Peng and Zhu, 2006; Xia et al., 2008b).

Shortcut nitrification=denitrification

Because NO�2 is an intermediate compound in both nitri-fication and denitrification, a pathway may be formed bycontrolling the extent of NHþ

4 oxidation to NO�2 (nitritation)instead of to NO�3 (nitratation) and then coupled by reductionof the accumulated NO�2 via denitrification (Ruiz et al., 2006;Yang et al., 2007).

The stoichiometric equations of the conventionalnitrification=denitrification process and the shortcutnitrification=denitrification process are shown below [Eqs.(1)–(6)] (Henze et al., 2008).

Conventional nitrification=denitrification:

Nitrification: NHþ4 þ 1:5 O2 ! NO�2 þH2Oþ 2Hþ (1)

NO�2 þ 0:5 O2 ! NO�3 (2)

Denitrification: NO�3 þ 4g CODþHþ

! 0:5 N2þ 1:5g biomass(3)

Summary: NHþ4 þ 2 O2þ 4g COD! 0:5 N2þH2O

þHþ þ 1:5g biomass(4)

Shortcut nitrification=denitrification:

Nitrification: NHþ4 þ 1:5 O2 ! NO�2 þH2Oþ 2Hþ (1)

Denitrification: NO�2 þ 2:4g CODþHþ

! 0:5 N2þ 0:9g biomass(5)

Summary: NHþ4 þ 1:5 O2þ 2:4g COD! 0:5 N2

þH2OþHþ þ 0:9g biomass(6)

From Equations (1)–(6), it can be seen that the shortcut ni-trification followed by denitrification of NO�2 can reduceapproximately 40% of carbon requirements, 25% of the oxy-gen supply, and result in 40% less biomass produced com-pared to oxidation to and denitrification of NO�3 . Thereductions in the COD requirement by this pathway makenitrogen removal from low C=N ratio domestic wastewaterfeasible. On top of that, NO�2 denitrification rates are 1.5–2times faster than NO�3 denitrification rates (Peng andZhu, 2006; Aslan and Dahab, 2008; Kornaros et al., 2008). Infull-scale WWTPs average maximum specific ammonia utili-zation rates (nitrification) are from 0.03 to 0.18 g-NHþ

4 -N=g-MLVSS � day (0.06–0.54 kg-NHþ

4 -N=m3=day), whereas theobserved specific denitrification rate (SDNR) are 0.04–0.42 g-NO�3 -N=g-MLVSS=day(0.08–1.36 kg-NO3

�-N=m3=day)for predenitrification with sufficient carbon source and0.01–0.04 g-NO�3 -N=g-MLVSS=day (0.02–0.12 kg-NO�3 -N=m3=day) for postdenitrification without an exogenous carbon(Kristensen et al., 1992; Munch et al., 1996; Tchobanoglouset al., 2004; Park et al., 2007).

112 SUN ET AL.

Unfortunately, NO�2 -N accumulation is difficult to attainbecause nitrite oxidizing bacteria (NOB) generally havehigher substrate-utilization rates than ammonia oxidizingbacteria (AOB) (Kim et al., 2008). However, it is technicallyfeasible to achieve shortcut nitrogen removal via NO�2 byspecific inhibition or removal of NOBs while retaining AOBs(Wu et al., 2007; Blackburne et al., 2008). These methods arebased on the fact that AOBs and NOBs have different physi-ological characteristics and responses to environmental fac-tors (Gali et al., 2006b). For instance:

1. Dissolved oxygen (DO) concentration. It has beenfound that NOBs have lower affinity for oxygen thanAOBs, with DO half-saturation coefficients of 1.2–1.5 mg=L and 0.2–0.4 mg=L, respectively. Therefore, thegrowth of NOBs can be limited by lowering the DOconcentrations at values between 0.3 and 1.2 mg-O2=L(Peng and Zhu, 2006; Chuang et al., 2007; Kim et al.,2008; Aslan et al., 2009).

2. pH value. Free ammonia (FA) and free nitrous acid(FNA) inhibit the growth of AOBs and NOBs at dif-ferent concentration levels. The concentrations of FAand FNA can be calculated according to Equations (7)and (8) as function of the pH value (Anthonisen et al.,1976),

FA(mg=L)¼ 17

14·

TAN · 10pH

[e6344=(273þT)þ 10pH](7)

FNA(mg=L)¼ 47

14·

TNN

[e�2300=(273þT) · 10pHþ 1](8)

where TAN is the total ammonium as nitrogen(TAN¼NHþ

4 - NþNH3 - N, mg - N=L), TNN is thetotal nitrite as nitrogen (TNN¼NO�2 - NþHNO2 - N,mg - N=L), T is the temperature (8C). Anthonisen et al.(1976) reported that from 0.1–1.0 mg-FA=L NOBs ac-tivity is inhibited, whereas concentrations between 10and 150 mg-FA=L are necessary to inhibit AOBs. In-hibition of NOBs by FNA is reported at concentrationsbetween 0.011 and 2.8 mg-FNA=L (Anthonisen et al.,1976; Vadivelu et al., 2006, 2007a, 2007b). Vadivelu et al.(2006, 2007a, 2007b), however, also reported thatgrowth of the AOBs (Nitrosomonas) become inhibited atapproximately 0.10 mg-FNA=L and are completelystopped at 0.40 mg-FNA=L. Although NOBs are in-hibited at concentrations as low as 0.1 to 1 mg-FA=L, thethreshold inhibitory concentration of FA graduallyincreases with time because NOBs can tolerate andacclimate to the increasing levels of FA (up to 40 mg-FA=L), resulting in an irreversible decline in NO�2 ac-cumulation efficiency (Wong-Chong and Loehr, 1978;Turk and Mavinic, 1989; Villaverde et al., 2000).

3. Temperature. It is possible for AOBs to effectivelyoutcompete NOBs at temperatures above room tem-perature (e.g., the specific growth rate of AOBs is ap-proximately two times that of NOB at 358C) (Bougardet al., 2006; Aslan and Dahab, 2008; Xue et al., 2009).Therefore, NOBs can be washed out in a contriuousstirred-tank reactor (CSTR) by controlling the sludgeretention time (SRT) at values between the critical SRT

for NOBs and AOBs. The SHARON (single reactorhigh-activity ammonia removal over nitrite) processrelies on successfully washing out NOBs from a reactoroperated at high temperature between 30 and 408C andan SRT of 1–1.5 days.

4. Inhibitors. Besides FA and FNA, some volatile fattyacids such as formic, acetic, propionic and n-butyricacid, and some heavy metals such as chromium (Cr),nickel (Ni), copper (Cu), zinc (Zn), lead (Pb), and cad-mium (Cd) preferentially inhibit NOBs (Peng and Zhu,2006). Eilersen et al. (1994) showed that concentrationsof formic acid above 100 mg=L inhibited NOBs withno adverse effect on AOBs. More recently, Blackburneet al. (2008) used 40 mg=L of formic acid to successfullyinitiate NO�2 accumulation in a sequencding batch re-actor (SBR) system. However, the use of inhibitors canalso bring an increase in the operation cost, the risk ofinhibiting other bacteria, and the possibility of themappearing in the effluent as secondary pollutants.

Table 1 summarizes the nitrogen removal from domesticwastewater based on shortcut nitrification=denitrification inrecent years. Total nitrogen (TN) removal rates between 0.08and 0.238 kg-N=m3=day were obtained for the domesticwastewater with a C=N ratio between 2.1 and 4.4 in a lab- andpilot-scale SBR reactor or A=O process.

Simultaneous nitrification=denitrification

Simultaneous nitrification=denitrification was inferred fromthe observation that TN losses of up to 30% are frequentlyreported in aerated tanks of the activated sludge process (Haoet al., 1997; Yoo et al., 1999; Zhao et al., 1999; Fuerhacker et al.,2000; Ammary and Radaideh, 2005). One prevailing explana-tion to this phenomenon is the simultaneous activity of auto-trophic nitrification on the surface of the activated sludge flocs,with denitrification occurring in the floc interior supported bythe existing DO gradients (Fig. 1) (Zhao et al., 1999; Holmanand Wareham, 2005). A second explanation is that the contin-ued ability of heterotrophic denitrifiers to reduce nitrogenunder aerobic conditions (Holman and Wareham, 2005). A lasttheory postulates that denitrifiers possess is greater physio-logical variety than previously expected. In particular, it hasbeen suggested that some denitrifiers are also heterotrophicnitrifiers and can directly convert NHþ4 into gaseous nitrogenend products (Zhao et al., 1999; Holman and Wareham, 2005).It should also be pointed out that AOBs such as Nitrosomonaseuropea and Nitrosomonas eutropha are able to denitrify in thepresence of small amounts of oxygen (Helmer and Kunst,1998), and some NOBs strains can denitrify as well (Schmidtet al., 2003). However, the contribution of these organismsto denitrification is speculative.

COD concentration, DO concentration, and floc sizeare significantly affecting the simultaneous nitrification=denitrification performance in activated sludge systems(Pochana and Keller, 1999). An optimal C=N ratio was re-ported at 11.1, where the nitrification and the denitrificationreactions are in balance (Chiu et al., 2007). The optimal bulkDO concentration was reported to be 0.5–1.5 mg=L (Bernatand Wojnowska-Baryla, 2007) with an optimum floc size of80–100 mm. Simultaneous nitrification=denitrification also oc-curs in some other wastewater treatment processes besidesthe activated sludge process, such as biofilm systems (Helmer

NITROGEN REMOVAL FOR LOW C/N DOMESTIC WASTEWATER 113

and Kunst, 1998; Hibiya et al., 2003), slow sand filters (Nakhlaand Farooq, 2003), biological aerated filters (Lemoine et al.,2006) and single packed bed batch reactors (Daniel et al.,2009).

Compared to the conventional nitrification=denitrificationpathway, simultaneous nitrification=denitrification offersseveral advantages (Andrade do Canto et al., 2008; Li et al.,2008b): (1) it utilizes 22–40% less carbon source and reducessludge yield by 30%, (2) it exerts less alkalinity demand be-cause alkalinity is consumed during nitrification but pro-duced during denitrification, (3) it has smaller footprintbecause it eliminates the need for either two separate tanksoperated in series or intermittent aeration operation in a sin-gle tank, (4) it consumes less energy due to the reductionin aeration requirement. Table 2 summarizes the differentprocesses for nitrogen removal from domestic wastewaterbased on simultaneous nitrification=denitrification. TN re-moval rates between 0.072 and 0.121 kg-N=m3=day havebeen achieved for domestic wastewater with a C=N ratio of2.7–4.9.

Anammox process

Anaerobic ammonium oxidation (Anammox) is an impor-tant microbial pathway in the nitrogen cycle that could beused to yield a new cost-effective technology to remove NHþ

4

from wastewater. Anammox bacteria (AnAOB) have the un-ique metabolic ability to oxidize NHþ

4 with NO�2 as electronacceptor to produce N2 and a small part of NO�3 under anoxicconditions without the requirement of an organic carbonsource (Date et al., 2008; Kuenen, 2008). AnAOBs belong to thePlanctomycetes. So far, two freshwater species, CandidatusBrocadia anammoxidans and Candidatus Kuenenia stuttgartien-sis, three marine species, Candidatus Scalindua sorokinii,Candidatus Scalindua brodae, and Candidatus Scalindua wag-neri, and a mixotrophic anammox bacterium CandidatusAnammoxglobus propionicus have been described (Tsushimaet al., 2007). The stoichiometry of the Anammox reaction isaccepted to be as in Equation (9) (Strous et al., 1998).

1 NHþ4 þ 1:32 NO�2 þ 0:066 HCO�3 þ 0:13 Hþ

! 1:02 N2þ 0:26 NO�3 þ 0:066 CH2O0:5N0:15

þ 2:03 H2O (9)

Compared to the conventional biological nitrogen removal,the Anammox process presents several advantages such as100% savings on external organic carbon source, 63% lessoxygen demand, small surplus biomass production, and re-duced carbon dioxide (CO2), nitrous oxide (N2O), and nitricoxide (NO) emissions (Feng et al., 2007; Liu et al., 2008).

Anammox has attracted a lot of attention in the wastewaterengineering field in recent years as a key element in alterna-tive nitrogen removal process. However, in order to takeadvantage of the Anammox, partial nitritation is required. Inthe oxidation of induced N-species (ammonia and organic-N),the conversion of NO�2 to NO�3 should be avoided and theamount of NHþ4 converted to NO�2 should be limited to ap-proximately 60% (Yamamoto et al., 2008). Many novel bio-logical nitrogen removal processes have been developed tocouple nitritation to Anammox. They can be grouped in twocategories: (1) two-reactor processes where a CSTR perform-ing partial nitrification is followed by another reactor per-forming the Anammox process (SHARON=Anammox) (VanDongen et al., 2001; Hwang et al., 2005), and (2) the one reactorprocesses with simultaneously nitritation and Anammox oc-curring, such as completely autotrophic nitrogen removalover nitrite (CANON) (Sliekers et al., 2002; Third et al., 2005)and oxygen-limited autotrophic nitrification denitrification(OLAND) (Windey et al., 2005). Both processes have beensuccessfully used for the treatment of concentrated NHþ4streams such as sewage sludge digester liquor, swine waste-water digester liquor, livestock wastewater, seafood processingwastewater, power plant wastewater, and other ammonium-rich wastewaters at lab-, pilot-, and=or full scales (Lamsamet al., 2008; Yamamoto et al., 2008). An outstanding case is thefirst full-scale Anammox reactor with a volume of 70 m3 andtreatment ability of 750 kg-N=day at the sludge treatment plantSluisjesdijk (Rotterdam, The Netherlands) (Van der Star et al.,

FIG. 1. Schematic diagramof simultaneous nitrification=denitrification in an activatedsludge floc.

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2007). Additional high TN removal rates of about 26 kg-N=m3=day have been obtained in an up-flow fixed-bed Ana-mmox biofilm column reactor (Tsushima et al., 2007).

Although Anammox can present a very cost-effective bio-logical nitrogen removal pathway, there are no report on thenitrogen removal from low C=N ratio domestic wastewaterby Anammox process. Some reasons for this maybe: (1) longreactor start-up periods due to extremely slow growth rates ofAnammox bacteria (with doubling time of about 11 days)(Strous et al., 1998). For the first full-scale Anammox reactor inRotterdam, no detectable Anammox activity was observedduring the first 800 days of operation, and stable conversiondid not take place until day 1,235 (Van der Star et al., 2007).Recently, a new cultivation system for Anammox bacteriawas reported in which the organisms grow in suspension atsignificantly higher enrichment, but the doubling time wasstill approximately 8 days (Van der Star et al., 2008). (2) Theirgrowth is reversibly inhibited by oxygen concentrationsabove 0.5% air saturation (about 0.042 mg-O2=L) and also ir-reversibly inhibited by NO�2 concentrations higher than70 mg-NO�2 -N=L (Schmidt et al., 2003; Kuenen, 2008). (3) Stilllittle is known about how the Anammox organisms interactwith denitrifiers, nitrifiers and dissimilatory nitrate reducerswhen organic compounds are present (Kuenen, 2008). BecauseAnAOBs have a much lower growth rate than denitrifiers, theyare easily outcompeted by denitrifiers, and Anammox reactioncannot be sustained if organic carbon (COD=N> 2–5) is pres-ent in the influent (Lackner et al., 2008). Therefore, moreexperience is necessary before Anammox is successfully in-corporated in the technologies for stable and reliable nitrogenremoval from low C=N domestic wastewater.

Figure 2 shows an indicative comparison between con-ventional nitrification=denitrification and the rest of the de-scribed microbial pathways. The reduction in organic carbondemand in the denitrification step, the lower oxygen re-quirements, and the reduced sludge production by these al-ternate pathways are the main differential features of theseprocesses. Additionally, higher removal rate can also be ob-tained by these pathways for low C=N ratio domestic

wastewater treatment compared to that by the conventionalnitrification=denitrification.

Alternative Novel Treatment Processes or Reactors

Common biological nitrogen removal processes occurs invarious treatment train configurations in WWTPs, includingthe A=O process, the A2=O process, the University of CapeTown (UCT) process, the Modified Bardenpho, and the Vir-ginia Initiative Plant (VIP) process. All these processes rely ona predenitrification zone where a portion of the nitrifiedwastewater is recycled and mixed with the influent to serve asan electron donor for denitrification. Disadvantages includethe need for high recirculation rates and the addition of ex-ternal carbon substrate when the influent C=N ratio is nothigh enough (Downing and Nerenberg, 2008b). To overcomethis situation, advanced process control methods, new bio-logical treatment processes and reactors such as the modifiedA2=O process, the multistage A=O step-feed process, mem-brane bioreactors (MBRs) and membrane-aerated biofilmreactors (MABRs) have been developed, many of them sup-porting the biological removal pathways described in previ-ous sections.

Advanced process control for achieving shortcutnitrogen removal

Control of DO, pH, and temperature have been success-fully used for shortcut nitrogen removal from high strengthammonium containing wastewaters such as sanitary landfillleachate, sewage sludge digester liquor, soybean wastewater,and industrial wastewater (Wang et al., 2004; Ciudad et al.,2005; Fux et al., 2006; Lai et al., 2004; Pedersen et al., 2003; Aslanand Dahab, 2008; Spagni and Marsili-Libelli, 2009). However,it is difficult and not economical to achieve shortcut nitrogenremoval for domestic wastewater treatment by maintaininga relatively high pH and temperature for several thousandor million tons=day of domestic wastewater in municipalWWTPs.

The online monitoring of parameters such as pH, DO,oxidation reduction potention (ORP), and N-species (NHþ

4 ,NO�2 and NO�3 ) can be helpful for achieving shortcut nitro-gen removal in domestic wastewater treatment process.Especially in batch mode, the end of nitrification and deni-trification can be readily obtained from the typical bendingpoints of pH, DO, and ORP profiles (Yang et al., 2007).Therefore, real-time control strategy regulating the operationtime of aerobic and anoxic phases can be successful toachieving shortcut biological nitrogen removal (Fuerhackeret al., 2000; Gao et al., 2003; Peng et al., 2004; Wu et al., 2007;Tanwar et al., 2008). In addition, studies have indicated thatexcessive energy and resource consumption could also beavoided as well by using a real-time control strategy.

For instance, Yang et al. (2007) and Wang et al. (2008b) usedthe real-time control strategy with pH (dpH=dt) and ORP(dORP=dt) as the control parameter in a pilot-scale SBR(working volume of 54 m3) and alternating activated sludgesystem (AAS) for real domestic wastewater treatment. Theresults showed that nitrogen removal via the nitrite pathwaycould be successfully achieved with an average NO�2 accu-mulation rate above 95% at either normal or low temperature(11.9–26.58C) in the SBR system and above 90% in the AAS

FIG. 2. Comparison the nitrogen removal by conventionalnitrification and denitrification, shortcut nitrification anddenitrification, simultaneous nitrification=denitrification, andAnammox process.

116 SUN ET AL.

system. In addition, approximately 20% aeration energy wassaved compared to the conventional steady-state controlsystems.

Modified A2=O process

The A2=O process with separate anaerobic, anoxic, andaerobic tanks is a suitable method for biological nitrogen re-moval (schema is shown in Fig. 3a). Denitrification of theNO�3 recirculated from a downstream aerobic tank occurs inan anoxic tank where denitrifiers can utilize the organicmatter present in the influent, avoiding the need for an ad-ditional organic carbon source (Ouyang et al., 2000). However,the A2=O configuration normally requires a high mixed liquidreturn ratio (2–4Qin) from the aerobic zone to the anoxic zoneto bring more NO�3 back for denitrification. High return ratioscan result in DO concentration increase and COD dilution inthe anoxic zone, which inevitably deteriorates the denitrifi-cation efficiency, especially when the organic matter presentin the influent wastewater is insufficient to deplete the DOpresent in the recycled mixed liquor. In addition, high returnratios also lead to higher energy consumption and increasedoperating costs.

Yang et al. (2006) modified the A2=O process to avoid thesedisadvantages (schema is shown in Fig. 3b). In the modified

A2=O process, (1) the recycle sludge is directed to a separatepreanoxic basin where hydrolysis processes can releasebiodegradable organic carbon, which can be used in the de-nitrification processes taking place downstream. (2) The in-fluent wastewater goes directly to the anaerobic zone of thereactor and gets mixed with the wastewater from the pre-anoxic tank. Part of mixed wastewater in this reactor is re-circulated to the postanoxic zone at a ratio of 0.4Qin to provideavailable organic carbon for denitrification. (3) In the firstaerobic zone, NHþ4 is oxidized to NO�2 and NO�3 , and bothare fed continuously to the so-called postanoxic zone for de-nitrification. (4) The treated wastewater passes through a finalaerobic tank to minimize the amount of COD in the effluent,and enhances the settling ability of the sludge by minimiz-ing denitrification in the secondary settler. Pilot-scale re-sults showed more than 88% COD and 70% TN was removedunder a volumetric loading rate of 0.9 kg-COD=m3=dayand 0.12 kg-N=m3=day for the real domestic wastewatertreatment.

Step-feed multistage A=O process

The step-feed multistage A=O process consists of two ormore denitrification–nitrification units in series with waste-water distributed at several reactor points so that an internal

FIG. 3. Schematic diagram of the anaerobic=anoxic=oxic (A2=O) process (a) and the modified A2=O process (b).

NITROGEN REMOVAL FOR LOW C/N DOMESTIC WASTEWATER 117

NO�3 recirculation is unnecessary. A schema of step-feedmultistage A=O process is shown in Fig. 4. The biodegradableorganic material in the influent is utilized for denitrification,and also simultaneous nitrification=denitrification may occurin this process (Zhu et al., 2007b).

For the design of a step-feed multistage A=O process it isimportant to optimize the influent flow distribution at eachstage, the relative split between anoxic and aerobic volumes,and the final pass mixed liquor suspended solid (MLSS)concentration (normally based on using an acceptable solidsloading for the secondary clarifier) (Tchobanoglous et al.,2004). The volume or sludge loading in each stage can sig-nificantly affect the nitrification rate. A small volume orsludge loading might lead to a decreasing of the nitrificationrate, whereas a large load might cause overload and sludgebulking (Zhu et al., 2007a, 2009).

Studies by Zhu et al. (2007a, 2009) indicated that a uniform lcan be used in this process for the treatment of wastewaterswith C=N ratio >9.0, by which it is simple for the processcontrol, optimization, and management. The optimal influentflow distribution ratio (lOpt) can be obtained by trial methodand the influent flow rate in each stage can be calculated byEquation (10).

QTotal¼ q � kn� 1Opt þ q � kn� 2

Opt þKþ q � kOptþ q (10)

where QTotal is the total influent flow rate (m3=day), q is theinfluent flow rate distributed in the last stage (m3=day), and nis the total number of the stage. However, the influent dis-tribution ratio (l) between two adjacent stages should not beuniform for treating domestic wastewaters with a low C=Nratio. An optimal l with 2.0:2.1:2.5:3.4 (0.2Qin:0.21Qin:0.25-Qin:0.34Qin) was proposed for the treatment of wastewaterwith C=N ratios lower than 4.9 in a step-feed four-stage A=Oprocess. Under the given condition, 20% higher TN removalefficiency could be achieved compared to the case with auniform l.

MBR

In an MBR, activated sludge treatment and membraneseparation are combined as a compact technology. MBRs have

several advantages over conventional activated sludge sys-tems including excellent removal of solids and organic matter,ability to maintain high biomass condition, ability to handle awide range of SRTs at relatively short hydraulic retentiontimes (HRT), and a small footprint due to the elimination ofclarifier (Ersu et al., 2008; Kimura et al., 2008).

In submerged MBR, good nitrification performance can beattained due to the high concentration of nitrifying biomass,long SRTs, and the aerobic condition provided by the inten-sive aeration used to mitigate membrane fouling. Almostcomplete nitrification can be easily achieved in MBR, butdenitrification may limit TN removal, because a requiredanoxic condition for denitrification cannot easily be createdwith intensive aeration (Kimura et al., 2008; Li et al., 2008b). Toovercome these drawbacks, some modifications of MBR havebeen proposed and developed:

1. A first approach involves installing an additional anoxicreactor with MBR for denitrification, typically in pre-denitrification configuration. Tan and Ng (2008) re-ported almost 0.1 kg-TN=m3=day was removed fromraw domestic wastewater with a mean C=N ratio of7.24 under mixed liquor recycle ratios of 3Qin using apredenitrification MBR.

2. A second approach involves creating cyclic anoxic andaerobic conditions in a single reactor MBR by inter-mittent aeration (Choi et al., 2008).

3. A final method, the MBR’s chamber is modified tocreate an anoxic or anaerobic zone within a single re-actor along with internal mixed liquor recirculation.

Chae et al. (2006) proposed a vertical submerged membranebioreactor (VSMBR) (Fig. 5a). In a VSMBR, lower anoxic zoneis separated from the upper aerobic zone by a horizontal platewith a hole in the center. The mixed liquor is recycled from theaerobic to the anoxic zone, while raw wastewater also entersthe anoxic zone for supporting denitrification by using in-fluent organic carbon. A pilot-scale study showed removalrates of 0.093 kg-TN=m3=day when treating domestic waste-water with a C=N ratio of 5.5 by the VSMBR (Chae and Shin,2007). Similarly, Kimura et al. (2008) proposed a baffledmembrane bioreactor (BMBR) (Fig. 5b). With a baffle insertedin the MBR’s chamber and an aeration bubbler placed inside

FIG. 4. Schematic diagram of the step-feed multistage anaerobic=oxic (A=O) process.

118 SUN ET AL.

of it, aerobic and anoxic conditions are alternatively created inthe outer zone at constant intervals. A pilot-scale studyshowed a 77% TN removal efficiency at a removal rate of0.102 kg-TN=m3=day for the treatment of domestic waste-water without external carbon addition.

Another possibility is to operate MBRs at very high MLSSlevels (25,000–30,000 mg=L), still allowing simultaneous nitri-fication and denitrification due to the creation of an inner ag-gregate anoxic zone to withstand heavy aeration (Sariogluet al., 2009). In addition, decay of the high concentration ofbiomass can support endogenous denitrification. By control-ling the bulk DO concentration at 1.5 mg-O2=L in a pilot MBRoperated without a separate anoxic volume, TN removal ratesof 0.19 kg-N=m3=day could be obtained for raw domesticwastewater with a mean C=N ratio of 7.1 (Sarioglu et al., 2009).Table 3 summarize the nitrogen removal results for the treat-ment of low C=N ratio domestic wastewater by various MBRsstudies.

MABR

The MABR is another novel bioreactor technology thatcombines biofilms and membrane technology. In an MABR,gas-permeable hydrophobic membranes are used for bub-bleless aeration and also as the support for biofilm growth.Through control of the intramembrane oxygen supply pres-sure and gas flow rate in MABR one can develop biofilms onthe membrane with oxic–anoxic–anaerobic zones (as shownin Fig. 6), which support stratification of different functionalmicrobial populations in a single biofilm (Li et al., 2008a;Syron and Casey, 2008a).

MABRs can support nitrification=denitrification biofilmswith higher activity compared to conventional biofilm reac-tors because of the unique biofilm stratification and high ox-ygen supply efficiency in MABRs. In a conventional biofilmreactor all substrates follow the same route of transportationand conversion processes within the biofilm. Because of theirhigher growth rates and oxygen affinities, heterotrophicbacteria generally outcompete autotrophs in the biofilm un-less the organic carbon concentration is low. This organiccarbon fluxes to the inner anoxic area are usually insufficientto support denitrification (Walter et al., 2005; Syron andCasey, 2008a). In an MABR, however, oxygen and nutrients

are provided from two opposite sides of the biofilm. Oxygenconcentration is highest at the biofilm–membrane interfaceand decreases toward the biofilm–liquid interface due to mi-crobial activity. At the same time the nutrient concentration ishigher at the biofilm–liquid boundary, decreasing with depthinto the biofilm. With such concentration gradients, popula-tions of nitrifiers and aerobic heterotrophs thrive adjacent tothe membrane while heterotrophic denitrifiers grow adjacentto the liquid because of the higher carbon concentrations andoxygen limitation there (Semmens et al., 2003; Satoh et al.,2004; LaPara et al., 2006; Lemoine et al., 2006; Downing andNerenberg, 2008a; Syron and Casey, 2008b). Therefore, anMABR can support complete nitrification=denitrification andcan offer a typically protected niche for nitrifiers againstheterotrophs, shock loads, and toxicants contained in thewastewater influent (Syron and Casey, 2008a).

Hibiya et al. (2003) reported simultaneous nitrification=denitrification in MABRs successful treating of domesticwastewater. By controlling the intramembrane oxygen supply,more than 80% TN and 90% TOC were removed under TNand TOC loading rates of 0.16 kg-N=m3=day and 0.42 kg-TOC=m3=day, respectively. FISH analysis showed that AOBswere mainly located inside the biofilm while denitrifyingbacteria developed in the outer biofilm layers and in the sus-pended sludge. Terada et al. (2003) reported 83% TN and 96%TOC removal efficiencies in an MABR under TOC and TNloading rates of 0.30 kg-TOC=m3=day and 0.27 kg-N=m3=day,respectively. Matsumoto et al. (2007) modeled the effects ofC=N ratio, biofilm thickness, and surface loading of oxygen onthe efficiency of MABR performing simultaneous nitrification=denitrification: by adjustment of the JO2

=JNHþ4

flux ratio couldcontrol simultaneous nitrification=denitrification and over70% TN removal efficiency could be achieved at C=N ratiosranging from 3.0 to 5.2 with biofilm thicknesses of 600–1,200mm. Hu et al. (2008) also reported that 86% COD, 94%NHþ4 and 84% TN removal in a carbon-membrane MABRunder COD and NHþ4 loading rates of 0.698 kg-COD=m3=dayand 0.154 kg-NHþ

4 -N=m3=day, respectively.In addition to traditional nitrification=denitrification,

completely autotrophic nitrogen removal can be achieved inMABRs. Terada et al. (2007) developed a multipopulationbiofilm model to compare the conventional and MABR bio-film reactor for autotrophic nitrogen removal: the MABR

FIG. 5. Schematic diagram of the modified membrane bioreactor (MBR): vertical submergedmembrane bioreactor (VSMBR)(a); baffled membrane bioreactor (BMBR) (b).

NITROGEN REMOVAL FOR LOW C/N DOMESTIC WASTEWATER 119

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biofilm had a wider range for autotrophic TN removal thanthat in conventional biofilm reactors and an optimal JO2

=JNHþ4

ratio of 1.73 g-O2=g-N was inferred. Furthermore, AnAOBactivity could be better maintained in the MABR biofilmversus conventional biofilm under excess oxygen conditions(JO2

=JNHþ4

ratio >3.7). Gong et al. (2007) developed a labora-tory-scale MABR where a gas-permeable carbon tube wascovered with nonwoven fabrics to support biofilm for single-stage autotrophic nitrogen removal. Partial nitrification andcompletely nitrogen removal were successfully achieved inthe single MABR, a max removal rate of 0.77 kg-TN=m3=daywas achieved with 88.5% TN removal efficiency after a 116-day startup without need for any external organic carbonsource addition. Moreover, FISH analysis showed that AOBs

and AnAOBs coexisted in the MABR biofilm. The nitrifyinglayer was located at the membrane–biofilm interface, whereasthe Anammox layer populated the anaerobic region closerto the liquid–biofilm interface (Gong et al., 2008). Clearly,MABRs are able to support single-stage completely autotro-phic nitrogen removal, offering a promising perspective fornitrogen removal from low C=N ratio wastewaters.

External Carbon Sources

Conventional carbon sources

Using the traditional nitrification=denitrification technol-ogy, external carbon sources are often required to supportdenitrification from nitrified low C=N domestic wastewater.

Table 4. Costs of Using Methanol, Ethanol, and Acetate for Postdenitrification

Carbon sourceCost

($=kg)C=N ratio

(g COD=gNO�3 - N)Yield of biomass

(g COD=g COD)Average SDNR

(gNO�3 - N=g VSS=day)Cost

($=Kg - NO�3 - N to N2)

Methanol 0.33 5.19 0.45 0.24 1.14Ethanol 0.84 6.08 0.53 0.71 2.43Acetate 1.03 8.40 0.66 0.74 8.15

FIG. 6. Schematic diagram of the simultaneous nitrification=denitrification (a) and completely autotrophic nitrogen removal(b) in membrane-aerated biofilm reactor (MABR) biofilm.

NITROGEN REMOVAL FOR LOW C/N DOMESTIC WASTEWATER 121

During the past 30 years, commercially available chemicalswith methanol, ethanol, acetic acid, propionic acid, and glu-cose have served to this end, with methanol typically pre-ferred because of its lower cost (Table 4) (Mokhayeri et al.,2009). Generally, only one type of carbon source is used, butrecent studies report that using mixed carbon sources mightbe more effective than using a single one. For instance, Hag-man et al. (2008) reported that a higher nitrate uptake rate wasobserved by the addition of a mixture of acetate and methanolthan with each alone. Ahmed et al. (2008) also reported thatusing methanol and acetate as a mixed carbon source wasmore effective for nutrient removal and fouling problemscontrol in an MBR. This can be explained by the presenceof two different populations and each metabolizing onlyone substrate, or a population of denitrifying bacteria thatutilized the two carbon sources simultaneously (Hagmanet al., 2008).

Nevertheless, the use of organic carbon sources increasesthe operating cost significantly, and less-costly alternativecarbon sources for denitrification have been explored in someWWTPs.

Alternative carbon sources

Various organic byproducts or waste rich in readily biode-gradable organic matter and volatile fatty acids (VFA), such aswine distillery effluent, the leachate of food waste, digestedpiggery manure, mecellulose wasted liquor, biodiesel waste,biologically hydrolyzed or mechanically disintegrated sludge,etc., have been suggested and tested as alternative carbonsources for denitrification (Table 5) (Kampas et al., 2007; Bodıket al., 2009). A potential benefit of using organic carbon con-tained in byproducts or waste materials as alternative carbonsources is that they are produced in large amounts and can beeasily available at zero or even at negative cost.

The preparation of these carbon sources involves acid oralkaline hydrolysis, thermal hydrolysis, thermochemical hy-drolysis, or biological hydrolysis to yield a concentrated and

complex hydrolysate containing low molecular weight andreadily biodegradable matter. Quan et al. (2005) used hydro-lyzed molasses as an external carbon source for nitrogen re-moval from synthetic wastewater in an SBR with a TNremoval efficiency of 91.6%, which was slightly superior tousing methanol (85%). Obaja et al. (2005) used digested andnondigested piggery manure as carbon source for biologicalnitrogen and phosphorus removal from digested piggerywastewater, same nutrient removal efficiencies (99.8% N and97.8% P) was also obtained as when using acetate.

Gali et al. (2006a) examined hydrolyzed primary and sec-ondary sludge as a carbon source for denitrification, but onlythe hydrolyzed primary sludge supported stable denitrifica-tion. Nitrogen removal efficiency was almost similar as whenusing methanol as a carbon source; it could lead to a costsaving of 0.2–0.3 euro for per kilogram-N removed. In addi-tion, Kampas et al. (2009) reported that mechanically disin-tegrated surplus activated sludge with energy inputs between2,300 and 6,200 kJ=kg-TS, and a short disintegration time(2–5 min) can also be used as an alternative carbon sourcefor denitrification.

However, it has to be pointed out that the addition of anyexternal organic carbon source to enhance nitrogen removalmay increase the load to the system, especially for using thealternative carbon sources that may also introduce slowlybiodegradable COD, leading to an increase of total suspendedsolids (TSS) in the process and compromise the effluentquality. Therefore, the dosage of external organic carbonsource addition should be prior tested and not exceed therequirement for the system.

Conclusions and Future Directions

Conclusions

The present work has reviewed innovative biological nitro-gen removal pathways, shortcut nitrification=denitrification,simultaneous nitrification=denitrification, and the Anammox

Table 5. Chemical Characteristics of Different Alternative Carbon Sources

Parameters

Pretreatedolive oil millwastewater

by lime(Tsonis, 1997)

Digestedpiggerymanure

(Obaja et al.,2005)

Nondigestedpiggerymanure(Obaja

et al., 2005)

Mecellulosewasted liquor(Park et al.,

2005)

Biodieselwaste(Bodık

et al., 2009)

Hydrolyzedprimarysludge

(Gali et al.,2006a)

Hydrolyzedsecondary

sludge(Gali et al.,

2006a)

Mechanicallydisintegrated

sludge(Kampas

et al., 2007)

VFA (mg=L) 1,050 5,275 2,000� 200 0 403–852BOD5 (mg=L) 1,730 3,250TCOD (mg=L) 68,200 3,085 7,450 58,846 1,600,000 60,000� 1,000 1900� 150SCOD (mg=L) 65,800 54,494 3,000� 200 55� 5 4,440–5,810NH4-N (mg=L) 1,600 720 180� 10 35� 5 50–60NO2-N (mg=L) 30� 3 <5NO3-N (mg=L) 20� 3 10� 1TN (mg=L) 720 19 2,000PO4-P (mg=L) 42 147 120 100–326TP (mg=L) 700TS (g=L) 11.21 13.42 35� 1 3.5� 0.05VS (g=L) 5.35 6.23 30� 1 2.2� 0.05TSS (g=L) 2.58 3.10 0.162 35� 1 2.2� 0.05VSS (g=L) 1.96 2.2 28� 1 2� 0.05

VFA, volatile fatty acids; BOD, biochemical oxygen demand; TP, total phosphorus; TCOD, total chemical oxygen demand; SCOD, soluablechemical oxygen demand; VSS, volatile susplended solids; TSS, total suspended solids; VS, volatile solids.

122 SUN ET AL.

process that can remove nitrogen with low or zero dosage oforganic carbon sources.

It can be concluded that the advanced process controlbased on the typical bending points of pH, DO and ORPprofiles is very helpful for achieving shortcut nitrogen re-moval in an activated sludge system. In addition, the modi-fied A2=O, step-feed mulistage A=O processes, MBR, andMABR, can support the innovative biological nitrogen re-moval pathways and they can be used to treat low C=N ratiosdomestic wastewater without external carbon addition.

Furthermore, it has been shown that alternative organiccarbon sources can be potentially used for enhanced biologi-cal nitrogen removal. These materials can offer the same orbetter performance for enhanced nitrogen removal comparedto the common carbon source, allowing for a substantial re-duction of the operating costs.

Overall, this review can provide guidance to municipalWWTPs operators and planners to select cost-effective nitro-gen removal alternatives for the treatment of low C=N ratiodomestic wastewater.

Future directions

Besides above, the presented results also show that furtherstudies are still required to obtain satisfying nitrogen removalfrom low C=N ratio domestic wastewater in WWTPs. Someresearch key points are proposed below: (1) go further andapply these new processes and technologies mentioned in thereview at a large scale for the treatment of low C=N ratiodomestic wastewater; (2) explore the best alternatives tointroduce the Anammox pathway into domestic wastewatertreatment process because of its attractive metabolism; (3)explore and optimize alternative carbon sources for enhancedbiological nitrogen removal. With these efforts, effective ni-trogen removal from low C=N ratio domestic wastewater canbe achieved in the near future.

Acknowledgments

The authors thank the Shanghai Tongji Gao Tingyao En-vironmental Science & Technology Development Foundation(STGEF) and the China Scholarship Council (CSC) for sup-porting this work. The authors also thank Lisbet Brusendorff(DTU Environment, Denmark) for her contributions to thegraph design in the present work.

Author Disclosure Statement

No competing financial interests exist.

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