1
Chapter 4 Principles of Biological Nitrogen Removal
2
1.0. SOURCES AND SINKS OF NITROGEN IN WASTEWATER TREATMENT
3
1.1
Influent Nitrogen Species
4
1.2
Ammonification
5
1.3
Ammonia Assimilation
6
1.4
Stripping
7
1.5
Denitrification
8
2.0 NITRIFICATION
9
2.1
Biochemistry and Microbiology
10
2.2
Stoichiometry
11
2.3
Kinetics
12
2.4
Toxicity
13
2.5
Biofilm Systems
14
3.0 DENITRIFICATION
15
3.1
Biochemistry and Microbiology
16
3.2
Stoichiometry
17
3.3
Substrate Requirements
18
3.4
Alkalinity Production
19
3.5
Kinetics
20
3.6
Toxicity
21
3.7
Biofilm Systems
22
4.0 REFERENCES
23
4-1
Final Draft
1
Nitrogen (N) is a key element involved in biogeochemical cycles. Its oxidation state can vary
2
from –3.0 for organic nitrogen to +5.0 for nitrate. The important nitrogen species are ammonium-
3
nitrogen (NH4+-N, -3), dinitrogen gas (N2, 0), nitrite-nitrogen (NO2--N, +3) and nitrate-nitrogen (NO3--
4
N, +5). Continuous interconversion between various nitrogen species in the environment is the basis of
5
the nitrogen cycle. The principal reactions involved in the nitrogen cycle are nitrogen fixation,
6
ammonification, nitrification, denitrification and assimilative nitrogen uptake. Because of the various
7
transformations that they undergo, the nitrogen species are rapidly converted from one oxidation state to
8
another. The various nitrogen species are summarized in Table 4.1.
9
10
11
[INSERT TABLE 4.1]
12
Although nitrogen is an essential component of building blocks of life, such as amino acids,
13
excessive deposition of nitrogen species such as ammonium-nitrogen (NH4+-N) and nitrate-nitrogen
14
(NO3--N) in terrestrial and oceanic ecosystems causes eutrophication and ground water pollution
15
(Robertson and Kuenen, 1992). Increasing industrialization has resulted in an increase in the global
16
nitrogen flux.
17
denitrification, is a greenhouse gas and, though not as abundant in the atmosphere as CO2, can
18
significantly contribute to global warming. The deleterious effects of NH4+-N include toxicity to aquatic
19
fauna, depletion of dissolved oxygen in receiving waters due to nitrification and a reduction in chlorine-
20
disinfection efficiency at water treatment facilities (U.S. EPA, 1993). Nitric acid (HNO3) can be a
21
principal contributor to acid rain. In addition to the well-documented damage to infrastructure, acid rain
22
can increase soil and water acidity. In fresh water, the nitrous and nitric acids produced by nitrification
23
can mobilize toxic aluminum ions, and can be lethal even by themselves to the flora and fauna
24
(Robertson and Kuenen, 1992). Further, the colonization of building surfaces by nitrifying
25
microorganisms can lead to corrosion and eventual destruction of structures due to the production of
26
nitrous and nitric acids. Yet another problem is the undesirable succession of the natural flora on
Nitrous oxide (N2O), which is a purported intermediate in both nitrification and
4-2
Final Draft
1
nutrient-poor land, when it is exposed to increased mineral nitrogen loadings. Nitrogen pollution of
2
environmental systems is caused by both anthropogenic and non-anthropogenic sources.
3
Because eutrophication and aquatic toxicity can result from the discharge of nitrogen containing
4
waters in aquatic environments, it is desirable to remove nitrogen species from wastewater before they
5
flow into sensitive water bodies. Though nitrogen removal from waters is possible using physico-
6
chemical methods (e.g. air-stripping at high pH and/or breakpoint chlorination for NH4+-N removal),
7
biological nitrogen removal (BNR) presents a more cost efficient and environmentally benign
8
alternative. The biochemical processes of nitrification and denitrification, acting in concert, channel
9
different nitrogen species through a part of the nitrogen cycle. The end product is dinitrogen (N2) gas,
10
which can be rapidly stripped from the BNR reactor. This chapter presents an overview of biological
11
nitrogen removal processes. A detailed discussion on nitrification and denitrfication is presented in
12
Chapter 5 and Chapter 6, respectively. In addition, sidestream nitrogen removal is presented in Chapter
13
9.
14
1.0. SOURCES AND SINKS OF NITROGEN IN WASTEWATER TREATMENT
15
1.1
16
Nitrogen species in typical domestic wastewater are predominantly in the reduced state with an
17
average oxidation state of –3, corresponding to that of ammonia (inorganic) or amino-acids (organic).
18
Based on an acid-base dissociation constant (pKa) value of 9.3 for the ammonia-ammonium pair
19
(Stumm and Morgan, 1996), most reduced inorganic nitrogen is in the form of protonated ammonium
20
ion (NH4+-N) rather than free ammonia (NH3).
21
collectively referred to and measured as total Kjeldahl nitrogen (TKN). TKN can be further classified
22
depending upon whether the reduced nitrogen species are soluble or particulate and biodegradable or
Influent Nitrogen Species
4-3
Ammonium/ammonia and organic nitrogen are
Final Draft
1
non-biodegradable. Free ammonia rather than ionic ammonium is the true substrate for nitrification
2
(Hooper et al., 1997).
3
4
1.2
5
Ammonification is a process by which organic reduced nitrogen (in the –3 oxidation state) is
6
converted to inorganic ammonium/ammonia. Ammonification is carried out by heterotrophic bacteria
7
and is a necessary pre-requisite to nitrification (Grady et al., 1999). The rate of ammonification is a
8
function of the organic nitrogen containing susbstrate concentration, the heterotrophic biomass
9
concentration catalyzing the ammonification reaction and the ratio of the carbon to reduced nitrogen
10
Ammonification
concentration of the wastestream (Grady et al., 1999).
11
12
1.3 Ammonia Assimilation
13
Ammonia-nitrogen (NH3-N) is the preferred assimilative nitrogen source for bacteria (nitrifying or
14
non-nitrifying) in activated sludge, since it is in the same oxidation state (-3) as in biomass
15
(approximated empirically by C5H7O2N) ( Hoover and Porges, 1952, Grady et al., 1999, Rittmann and
16
McCarty, 2001). This means that bacteria do not need to reduce ammonia, unlike all other more
17
oxidized nitrogen species such as dinitrogen gas, nitrite or nitrate (Grady et al., 1999, Rittmann and
18
McCarty, 2001)). The relative nitrogen content of biomass and correspondingly the amount of ammonia
19
assimilated is 0.0875 g-N/g of particulate COD formed (Grady et al., 1999). In the absence of ammonia
20
or reduced organic nitrogen (assimilated after ammonification), activated sludge can assimilate more
21
oxidized sources such as nitrite or nitrate, but at a significant expenditure of energy (or electron
22
equivalents) required to reduce these species to the –3 oxidation state (Rittmann and McCarty, 2001).
23
The growth of microorganisms or biomass represented by C5H7O2N can be shown in a simplified
24
manner as follows:
4-4
Final Draft
1
2
Organic matter + O2 → CO2 + C5H7O2N (new biomass) + energy + other products
(4.1)
3
4
5
Biomass present in the wastewater treatment system is oxidized through endogenous respiration,
which can be represented by:
6
7
C5H7O2N (biomass) + 5 O2 → 5 CO2 + NH3 + 2 H2O + energy + other products
(4.2)
8
9
10
Thus endogenous respiration releases some of the nutrients back in to the wastewater treatment
process.
11
12
1.4
13
Non-ionized free ammonia, NH3 is significantly volatile and could be potentially removed by
14
stripping from the activated sludge in the aerated zones. However, typical activated sludge plants
15
operate at a pH of close to 7.0. At this pH value, the liquid phase NH3 concentration is more than two-
16
orders of magnitude lower than the ionized (non volatile) ammonium NH4+ form. Thus, stripping is
17
expected to contribute minimally to the overall ammonia removal from activated sludge trains.
Stripping
18
19
1.5 Denitrification
20
In addition to removal of nitrogen through incorporation into new biomass growth and limited
21
stripping that might occur under suitable conditions, the other major route of nitrogen removal from
22
wastewater treatment processes is the biological reduction of nitrate and nitrite to primarily nitrogen gas,
23
N2 in the denitrification reaction. The atmosphere acts as a nitrogen sink where nitrogen in gaseous form
24
is the principal form of nitrogen.
4-5
Final Draft
1
The overall amount of total nitrogen removed through the system depends upon the amount of WAS
2
generated and the dentrification occurring in the process. The amount of WAS, in turns, depends upon
3
the solids retention time (SRT) of the activated sludge process.
4
5
6
7
2.0 NITRIFICATION
Details of nitrification in wastewater treatment are given in Chapter 5. However, the following
sections provide an overview of the nitrification process.
8
9
10
11
2.1 Biochemistry and Microbiology
12
Nitrification is the process of biological oxidation of ammonia (which exists mostly as NH4+-N in
13
typical wastewater) to nitrite (NO2–-N) and further oxidation of nitrite to nitrate (NO3–-N). The oxidation
14
of ammonia to nitrite is carried out by ammonia oxidizing bacteria (AOB) and nitrite conversion to
15
nitrate is carried out by nitrite oxidizing bacteria (NOB). Both of these reactions should operate at
16
optimal rates for production of nitrate. Ammonia and nitrite oxidizers are collectively referred to as
17
nitrifiers. Although classified together, AOB and NOB are not related phylogentically (Bock et al.,
18
1991).
19
Most nitrifiers found in typical wastewater treatment systems are autotrophic, as they synthesize
20
cellular material from inorganic carbon (HCO3–) under typical operating conditions. Oxidation of the
21
ammonia or nitrite provides the energy needed for cell synthesis. These bacteria are obligate aerobes, as
22
they grow only when dissolved oxygen (DO) is available. The absence of DO for prolonged periods,
23
however, is not lethal (Painter, 1970) as these organisms adapt and survive under low DO as well as low
24
ammonia concentrations (Geets et al., 2006). In typical biological nutrient removal (BNR) systems,
25
nitrifiers successfully survive anaerobic (absence of oxygen and oxidized nitrogen species) and anoxic
26
(absence of oxygen but the presence of oxidized nitrogen species) conditions, which they successfully
27
do. The relative abundance and diversity of nitrifying organisms in wastewater treatment systems
4-6
Final Draft
1
depends on influent characteristics and operating conditions (Siripong and Rittman, 2007, Ahn et al.,
2
2008).
3
Heterotrophic nitrification has been reported by some researchers (Joo et al., 2005, Su et al., 2006,
4
Lin et al., 2006, Joo et al. 2007). The microorganisms capable of heterotrophic nitrification include
5
Thiosphaera pantotropha, Bacillus, Pseudomonas, Alcaligenes, Pseudomonas denitrificans, and
6
Paracoccus denitirficans. These bacteria generally carry out aerobic denitrification as well and
7
contribute to simultaneous nitrification and denitrification. The heterotrophic nitrification process
8
requires a readily available organic substrate, such as acetate, which is usually very limited in aerobic
9
zones. As a result, heterotrophic nitrifier population is likely to be insignificant in most municipal
10
wastewater treatment plants (van Loosdrecht and Jetten, 1998). Ammonia can be utilized as an inorganic
11
electron donor in the presence as well as absence of oxygen. In the absence of oxygen the reaction
12
occurs with nitrite as the electron acceptor in the anammox (anaerobic ammonia oxidation) process.
13
Electrons from ammonium are transferred to nitrite producing nitrogen gas and water
14
(www.anammox.com; Strous et al., 1999; Egli, 2003; Egli et al., 2001):
15
16
NH4+ + NO2– → N2 + 2 H2O
(4.3)
17
18
The anammox process adds a significant amount of gasoues nitrogen to the atmosphere (Arp and
19
Bottomley, 2006). As of now anammox bacteria have not been obtained in pure culture. The bacteria
20
capable of anammox include Candidatus Brocadia anammoxidans, Candidatus Kuenenia stuttgartiensis,
21
Candidatus Scalindua brodae, Candidatus Scalindua wagneri, and Candidatus Scalindua sorokinii (Egli
22
et al., 2001; Schmid et al., 2000; Schmid et al., 2003; Kuypers et al. 2003). These organisms are
23
classified under phylum Planctomycetes, class Planctomycetacia, and order Planctomycetales (Perxas,
24
2005). Anammox process is very slow, as the doubling time of the anammox bacteria is 10.6 days with
25
the maximum specific growth rate estimated to be 0.003/h (Jetten et al., 2001). Recent research,
4-7
Final Draft
1
however, indicates that by providing suitable seed biomass and reactor operating conditions, it is
2
feasible to oxidize ammnonia at very high rates using the anammox process (Tsushima et al., 2007). The
3
main advantages of the anammox process include: (a) significant reduction in nitrification oxygen
4
requirement, because only half of the influent nitrogen needs to be nitrified, and only to NO2–-N, (b)
5
reduction in COD or supplemental organic carbon requirement for denitrfication, (c) substantial
6
reduction in production of WAS, and (d) elimination of CO2 emissions that occur during conventional
7
denitrification.
8
To overcome the slow growth and consequent washout of nitrifiers from high-rate systems, especially
9
at low temperatures, a process named “In-Nitri®” (Inexpensive Nitrification) has been proposed (U.S.
10
EPA, 2007a). The In-Nitri® process consists of a nitrifying system to exclusively treat side-streams in
11
treatment plants. Nitrifiers are grown using ammonia from digested sludge, sludge dewatering liquid, or a
12
commercial source. The WAS from this system is fed into the mainstream aeration tank, where NH4+-N
13
can be nitrified by the constant supply of nitrifiers. In-Nitri® process is operated at short SRT and takes
14
advantage of higher ammonia concentrations and temperatures in side-streams, for example, treatment
15
of warm anaerobic digester supernatant having 30-35°C.
16
SHARON (single-reactor high-activity ammonia removal over nitrite), CANON (completely
17
autotrophic nitrogen removal over nitrite), and OLAND (oxygen-limited autotrophic nitrificaion-
18
denitrification) are some of the other newer processes developed to oxidize ammonia. In the SHARON
19
process, ammonia is is oxidized to nitrite rather than nitrate and denitrified, which saves 25% of the
20
oxygen requirement for nitrification and 40% of the external carbon (as methanol) in denitrification, as
21
shown in Equations 4.4 and 4.5 (Jung et al., 2007). The key to success of the SHARON process is the
22
elimination of NOB from the system by operating at low SRTs (1-2 d) and high temperatures (>25ºC),
23
where AOB out-compete NOB (Van Hulle et al., 2005; Paredes et al., 2007):
24
25
NH4+ + 1.5 O2 + 2 HCO3– → NO2– + 2 CO2 + 3 H2O
(4.4)
4-8
Final Draft
1
2
6 NO2– + 3 CH3OH + 3 CO2 → 3 N2 + 6 HCO3– + 3 H2O
(4.5)
3
The CANON process is completely autotrophic and does not require organic carbon for
4
denitrfication (Third et al., 2001). In this process, ammonium is first converted to nitrite under aerobic
5
conditions and then the nitrite is converted to nitrogen gas in the absence of oxygen (Equation 4.6). In
6
the OLAND process, AOB are able to convert ammonia to nitrogen gas under oxygen-limited conditions
7
in a single reactor (Equation 4.7) (Kuai and Verstraete, 1998).
8
9
NH3 + 0.85O2 Æ 0.11NO3- + 0.44N2 + 0.14H+ + 1.43H2O
(4.6)
2NH4+ + 1.5O2 → N2 + 3H2O + 2H+
(4.7)
10
11
12
13
Anammox, SHARON, CANON, and OLAND processes are suitable for treating wastewaters with
14
high ammonia concentrations (e.g. leachate from landfills, anaerobic digestion dewatering liquors),
15
usually operating at warm temperature (>25oC). The flow rate of supernatant from digesters is relatively
16
small, however, it can can contribute to 10-20% of the total nitrogen load to the treatment plant in a
17
BNR system. In other words, in a conventionally operated activated sludge plant (without BNR), the
18
recycle from the digesters can increase effluent ammonia-nitrogen by 50% or more. This stream is an
19
ideal source to implement novel nitrogen removal processes. Additional details on sidestream nitrogen
20
removal are presented in Chapter 9.
21
22
2.2 Stoichiometry
23
The first step of ammonia-oxidation to nitrite-nitrogen in the overall nitrification process can be
24
written as follows:
25
4-9
Final Draft
1
NH4+ + 1.5 O2 → NO2– + 2H+ + H2O
(4.8)
2
3
4
This is an energy reaction and does not include the production of cell mass. The nitrite produced is
then oxidized as follows:
5
6
NO2– + 0.5 O2 → NO3–
(4.9)
7
8
The overall energy reaction can be written by summing the above two equations:
9
10
NH4+ + 2 O2 → NO3– + 2 H+ + H2O
(4.10)
11
12
Based on the stoichiometry of the overall energy reaction, 2 moles of oxygen are required to oxidize 1
13
mole of nitrogen to nitrate, which is equivalent to 4.57 g O2 per g NH4+-N oxidized. It should be noted
14
that these reactions do not take biosynthesis into account and additional equations that include
15
biosynthesis are provided in Chapter 5.
16
17
2.3 Kinetics
18
Ammonia oxidizing bacteria (AOB) typically grow slower than nitrite oxidizing bactera (NOB),
19
making the first step in nitrification as the rate limiting step. In the past, kinetics of nitrification were
20
modeled in a single step (NH4+-N → NO3–-N). However, recent research suggests that this is not
21
appropriate and modeling of individual oxidation reactions in two-steps is necessary (Chandran and
22
Smets, 2000). At higher temperatures (25-35ºC) and low SRTs (1-2 days), AOB indeed grow faster than
23
NOB. As a result, NOB are washed out of the system and only the first step of nitrification can be
24
accomplished, which is exploited in the SHARON process. .
25
4-10
Final Draft
1
2
2.4 Toxicity
3
Nitrifiers are less robust than heterotrophs and their performance is sensitive to a number of heavy
4
metals and synthetic organic chemicals, as summarized in Table 4.2. The effects of the substances can be
5
either inhibitory or fatal depending on the compound, its concentration, the duration of exposure, and
6
other environmental conditions in the nitrification reactor. Special mention should be made of the
7
inhibitory effect of gaseous (free or un-ionized) ammonia [NH3(g)] and un-ionized nitrous acid (HNO2).
8
Nitrosomonas and Nitrobacter are inhibited by free ammonia (150 mg/L) and nitrous acid (Anthonisen
9
et al., 1976; Turk and Mavinic, 1986). Table 4.3 summarizes the range of ammonium and nitrite
10
concentrations that may inhibit nitrification. Nitrobacter appears to be more sensitive to free ammonia
11
than Nitrosomonas. Turk and Mavinic (1986) found that nitrite oxidation was inhibited at 0.1-1.0 mg/L
12
NH3(g)-N (free ammonia) and that ammonia oxidation was inhibited at 5-20.0 mg/L NH3(g)-N. Ford et
13
al. (1980) found that nitrite oxidation was inhibited at 10-150 mg/L NH3(g)-N.
14
At low pH, ammonia oxidation is more sensitive to nitrite than nitrite oxidation. Inhibition by nitrite
15
at low pH is likely caused by the presence of free nitrous acid (Beccari et al., 1979). Turk and Mavinic
16
(1986) also found that nitrite-nitrogen levels as high as 100 mg/L caused no discernible inhibition to the
17
treatment process. The conclusion again is that free nitrous acid, rather than nitrite, is likely inhibiting
18
the process.
19
Alleman (1984) observed elevated nitrite concentration of 27 mg/L in batch nitrification systems and
20
concluded that Nitrobacter is more susceptible than Nitrosomonas to environmental stresses. Alleman
21
(1984) proposed that reduced temperature, limited oxygen and carbon dioxide, elevated pH, presence of
22
free ammonia, and excess solids wasting reduce Nitrobacter growth and nitrite oxidation. Alleman
23
(1984) reported that shock ammonium loading and reduction of nitrate induce nitrite accumulation.
24
Shock loads of ammonium can cause nitrite accumulation because Nitrosomonas can adapt its
25
population more quickly than Nitrobacter. Mines (1983) observed nitrite concentrations ranging from 19
4-11
Final Draft
1
to 210 mg N/L at free ammonia concentration from 9.5 to 73 mg/L in continuous-flow studies treating
2
high-strength nitrogenous wastewater. Tanaka and Dunn (1982) found that oxygen concentrations
3
approaching zero led to nitrite concentrations of 36 mg/L in a laboratory-scale fixed-film batch
4
nitrification system. They also found that when the DO concentrations increased, the nitrite levels de-
5
creased to below 10.0 mg/L.
6
Salinity can impact nitrification rates. Panswad and Anan (1999) reported decrease in specific
7
ammonia-nitrogen uptake rate of nitrifying sludge acclimated to varying levels of NaCl. They operated
8
BNR systems acclimated to 0-30 g/L of NaCl. After dosing them with 70 g/L of NaCl for 4 days, they
9
observed the specific ammonia-N uptake rate (SAUR). For sludge not acclimated to NaCl, SAUR
10
dropped from 4.76 to 0.48 mg NH4+-N/g MLSS.h within 3 days. After recovery, SAUR came back up to
11
4.33 mg NH4+-N/g MLSS.h. Sludge acclimated to 30 g/L of NaCl showed a decrease in SAUR from
12
2.14 to 1.02 mg NH4+-N/g MLSS.h but recovered to 2.13 mg NH4+-N/g MLSS.h. Sludge acclimated to 5
13
and 10 g/L NaCl recovered to better SAURs than before the shock, suggesting that nitrifiers are affected
14
by salinity, however, they are also capable of recovering from the salinity shock. A summary of salt
15
effects on nitrification is presented in Table 4.4 (Paredes et al., 2007).
16
17
18
19
Inorganic substances such as cadmium, chromium, cyanide, arsenic, fluoride, nickel, and zinc can lead
20
to nitrification inhibition (Hu et al., 2003; Hu et al., 2002; Fox et al. 2006). Fox et al. (2006) concluded
21
that significant inhibition occurred at 10 mg/L of Zn, whereas, complete inhibition occurred at 50 mg/L
22
of Zn. In the same study, only a slight inhibition was observed at 1 mg/L of Zn. Hu et al., (2002) found
23
that free cation concentration of nickel and cadmium, and not the total aqueous concentration of the
24
metal, correlated with nitrification inhibition. In addition, it was observed that the addition of chelating
25
agent such as EDTA reduced the inhibitory effect. Nevertheless, excessive use of chelating agent itself
26
can be inhibitory. Therefore, care needs to be exercised in the selction of chelating agent concentration.
[INSERT TABLES 4.2, 4.3, 4.4]
4-12
Final Draft
1
In general, the free ion concentration in activated sludge processes is reduced due to production of
2
exocelluar polymers, which bind some of the metals, reducing the overall inhibition capacity of the
3
metals. Hu et al (2003) investigated the impact of copper, cadmium, nickel and zinc on nitrification. The
4
objective of this study were to (i) evaluate the relationship between metal partitioning and metal
5
inhibition for a mixed nitrifying consortium; (ii) determine metal internalization kinetics after transient
6
exposure of metals; and (iii) develop a mathematical model to describe nitrification inhibition that
7
captures both metal transport and biological toxicity effects. The results of this study indicated that, in
8
short-term batch assays (about 1 h), the specific ammonium oxidation rate decreased as the applied
9
metal dose to nitrifying biomass increased for all metals. The metal molar inhibitory effect toward
10
ammonium oxidation was cation specific and followed Cu2+ Zn2+ > Cd2+ > Ni2+. It was also observed
11
that the nitrification inhibition increased with exposure time; however, sorbed metal concentrations were
12
not good predictors of the metals’ effect on nitrification kinetics. On the contrary, inhibition of
13
ammonium oxidation kinetics correlated well with the intracellular Zn, Ni, and Cd concentrations,
14
indicating that intracellular Zn, Ni, and Cd were directly responsible However, the results for copper
15
showed no direct correlation between intracellular or sorbed Cu concentrations and nitrification
16
inhibition. Copper is characterized by high complexation potential, high degree of partitioning to
17
nitrifying biomass, and fast internalization kinetics. It was suggested that the difference in
18
physicochemical behavior of copper compared to other metals studied appeared to have different
19
biological response. The precise method of copper inhibition on nitrfication was not clearly understood
20
and further research is needed to better undertsnd the mode of copper toxicity or inhibition to ammonia
21
oxidation. .
22
23
2.5 Biofilm Systems
24
In addition to suspended-growth systems, a variety of attached-growth or fixed-film systems are
25
applied to nitrify domestic and industrial wastewater. In these systems, biomass is attached to solid
4-13
Final Draft
1
support media contained within a reaction vessel. The wastewater to be treated is brought in contact with
2
the biofilm, where the local mixing and turbulence determines the transfer of nutrients to the biofilm.
3
The growth of biofilm needs to be balanced from excessive detachment to avoid clogging of the reactor,
4
while maintaining activity within the bioreactor (Henze et al., 2008). As in the case of suspended growth
5
systems, addition of limiting nutrients, alkalinity, etc is possible to maintain the desired performance of
6
the reactor. Although the kinetic (growth and oxidation) relationships used in the design of suspended-
7
growth reactors are still valid, the design of attached-growth systems is complicated by mass-transfer
8
limitations of substrate within the biofilm system. The biofilm system can be described by four
9
components consisting of bulk liquid, boundary layer, biofilm, and substratum, as shown in Figure 4.1
10
(Eberl et al., 2006.). In the past, the design of fixed-film processes was mainly based on empirical data
11
derived from pilot or full-scale system, however, recent advances in biofilm modeling has made these
12
models available in commercial process simulators to facilitate design of fixed-film processes (Eberl et
13
al., 2006). Many different types of media (shape, size, and material of construction) are commercially
14
available and the choice depends on the reactor design. Factors that should be considered in the selection
15
of media include: specific surface area, density, material of construction, attrition resistance, and
16
suitability for biofilm attachment (Lazarova and Manem, 2000). The larger media have more void
17
spaces and reduced risk of clogging but have lower specific surface area, increasing the overall size of
18
the reactor volume. Whereas, smaller size media have opposite characteristics of higher specific surface
19
area, higher risk of clogging and smaller reactor volume. One has to strike a balance between the
20
potential for clogging and reactor size for a given application to optimize the design. Other design
21
features that should be be considered in the design of fixed-film reactors include: aeration, flow
22
distribution, biofilm control and solids removal.
23
In fixed-film nitrification reactors, the competition or interactions between heterotrophic and
24
autotrophic bacteria may be more important than in suspended-growth systems. The presence of
25
significant amount of organic substrate in the fixed-film reactor can allow the heterotrophic biomass to
4-14
Final Draft
1
overwhelm the autotrophs and effectively prevent their growth until the carbon substrate concentration
2
in the bulk liquid is reduced to approximately 20 mg/L or less soluble biochemical oxygen demand
3
(sBOD5) (Boller et al., 1994). An overview of the important parameters that impact nitrification was
4
presented by Boller et al. (1994).
5
6
[INSERT FIGURE 4.1]
7
8
Mass transfer of nutrients and competition for DO between heterotrophic and autotrophic bacteria
9
become more critical in attached-growth systems. The concentration of substrates such as ammonia-
10
nitrogen and DO within the biofilm and external liquid layer can be significantly lower than in the bulk
11
liquid due to transport limitations. Low concentrations within the biofilm can result in lower rates of
12
nitrification. To accurately describe the processes occurring in biofilms, diffusion-reaction models that
13
consider both external (liquid film) and internal (biofilm) mass-transfer resistances need to be
14
constructed. Previously, it was understood that the external (liquid film) mass-transfer resistance is
15
negligible in comparison to the internal resistance. However, more recent work suggests that the mass-
16
transfer within the external liquid layer or the boundary shown in Figure 4.1 is equally important to the
17
process (Eberl et al., 2006).
18
Traditional examples of attached-growth systems include trickling filters, rotating biological
19
contactors, submerged packed-bed reactors, and fluidized bed reactors. Immobilized cells or high
20
biomass processes also rely significantly on attached-growth biomass. Trickling filters are one of the
21
oldest type of biofilm reactors, where 5-20 cm large rocks or plastic material as static support media.
22
The height of the tricking filter may range from 1-3 m for rock meadi and 4 to 12 m when using plastic
23
media (Henze et al., 2008). Typically, attached-growth systems tend to be more resistant to shock loads
24
and, when the reactors are enclosed, they are also protected from excess loss in temperature aiding the
25
process. An additional advantage of attached-growth systems is that there is no need for sludge
4-15
Final Draft
1
recirculation to maintain the necessary biomass for treatment because the biomass is attached to the solid
2
support in the reactor. From a practical design viewpoint, this means process efficiency is not dependent
3
on the settleability of the biomass, unlike the necessity for suspended-growth systems. However, short
4
detention time, in some of the fixed-film reactors may result in breakthrough of ammonium-nitrogen at
5
peak flows. As with suspended-growth systems, nitrification may be accomplished in a separate unit
6
process or in combination with carbonaceous removal in a single reactor.
7
During the past decade, the effluent ammonia-nitrogen limits have become more stringent. This has led
8
to upgrade of a number of existing plants with the addition of synthetic media to increase nitrification
9
capacity of the system. These integrated fixed-film activated sludge (IFAS) systems are hybrid reactors
10
that use synthetic media completely submerged into water (either fixed in the aeration tank or suspended
11
in the mixed liquor) and sludge is recycled from the secondary clarifier. When sludge is not recycled,
12
which is typically the case for separate stage nitrification processes, suspended media are employed in
13
moving bed biological reactors (MBBRs). These systems allow growth and retention of additional
14
biomass in the reactor without the need for an increase in clarification capacity of the system. The
15
additional biomass results in higher SRT, improving carbon removal and nitrification rates. A similar
16
concept is used in a biological aerated filter (BAF). The BAF types of system provides both biological
17
treatment as well as filtration of solids. In a submerged aerated filter, rigid, corrugated, structured
18
polypropylene media are installed in an aeration tank to provide a high surface area for biomass
19
attachment (500-1150 m2/m3 or 150-350 ft2 per ft3). The media are arranged into cells-in-series in which
20
effluent is contacted with the fully submerged media in the presence of co-current aeration
21
(www.severntrentservices.com).
22
23
24
25
26
3.0 DENITRIFICATION
Denitrification or reduction of nitrate to nitrogen gas under anoxic conditions depends on nitrate
being produced in the nitrification process under aerobic conditions. Therefore, for the total nitrogen
4-16
Final Draft
1
removal, first nitrification and then denitrification should occur efficiently to achieve the desired effluent
2
quality. Nitrification requires aerobic conditions and consumes alkalinity, whereas, denitrfication does
3
not require aerobic conditions and generates NO3-N as the alternate electron acceptor, which reduces the
4
overall oxygen requirement of the process. Denitrfication also returns part of the alkalinity consumed
5
during nitrification. Thus, where feasible, denitrfication should be incorporated to reduce the total
6
energy footprint and the external alkalinity addition to the process. The potential disadvantage is the cost
7
of adding external carbon (e.g. methanol) when wastewater does not contain sufficient amount of readily
8
biodegradanble carbon to meet the effluent total nitrogen limits.
9
10
3.1 Biochemistry and Microbiology
11
Most denitrifiers are facultative, meaning they can use either oxygen or oxidized nitrogen (NO2–-N or
12
NO3–-N) as the terminal electron acceptor in respiration. The use of oxygen as the electron acceptor is
13
termed as aerobic respiration, whereas, the use of nitrate or nitrite as electron acceptor is termed as
14
anoxic respiration. These microorganisms use similar metabolic pathways. A major difference between
15
aerobic respiration and anoxic respiration is the enzyme catalyzing the final electron transfer occurring
16
in the electron transport chain. Oxygen must be excluded to promote dissimilatory denitrification, which
17
is the process in which nitrate is used an an alternative electron acceptor (Madigan et al., 1997). If both
18
oxygen and nitrate are present, microorganisms preferentially use oxygen as the terminal electron
19
acceptor, because it yields more energy than nitrate or nitrite. The advantages of removing wastewater
20
COD through denitrification include: (a) reduction in aeration requirement for the process, (b) a slight
21
reduction in the overall sludge production as biomass yield in anoxic conditions is less than the yield in
22
aerobic conditions, (c) recovery of alkalinity, (d) effluent with low nitrates, which reduces negative
23
impact on the receiving water, and (e) a reduction in filaments thus better settling solids.
24
Microorganisms require nitrogen for protein synthesis. The preferred source of nitrogen is NH4+-N
25
because this form is used directly in synthesis. Nevertheless, if sufficient NH4+-N is unavailable, some
4-17
Final Draft
1
microorganisms can reduce nitrate to ammonium (Gayle et al., 1989). This process is referred to as
2
assimilatory nitrate reduction (NO3– → NO2– → NH2OH → Organic nitrogen), indicating that nitrogen
3
is incorporated into the cell. This reaction can proceed successfully even under aerobic conditions
4
(Madigan et al., 1997). It is therefore distinguished from dissimilatory nitrate reduction (denitrification),
5
which is a respiratory process whereby the microorganism obtains energy. Four steps are involved in
6
dissimilatory biological denitrification (Grady and Lim, 1980):
7
8
NO3– → NO2– → NO (g) → N2O (g) → N2 (g)
(4.11)
9
10
NO2–, NO, and N2O are intermediates in the process. Each step involves a particular reductase
11
enzyme that catalyzes the transfer of electrons to nitrogen. Nitrate reductase, a molybdenum-containing
12
enzyme converts NO3– to NO2– and nitrite reductase catalyzes the conversion of nitrite NO2– to NO.
13
Nitric oxide reductase converts NO to N2O, and in the final step, nitrous oxide reductase produces
14
gasesous nitrogen. NO and N2O are both non-ionic gaseous forms of nitrogen. N2O is especially
15
important in that it is a significant greenhouse gas released from wastewater treatment.
16
The electrons originate from the substrate, that is, the electron donor. Either inorganic (for example,
17
hydrogen or sulfur) or organic waste compounds can serve as substrate for denitrification. As a result of
18
denitrification, the electron donor is oxidized while nitrate is reduced. In addition to organic material
19
present in the wastewater, external carbon sources are frequently used to provide a source of electron
20
donor for denitrfication. The possible electron transport system of the dissimilatory denitrification is:
21
22
e- donor → NAD → FAD → Quinone → Cytochrome → Nitratereductase → NO3–
(4.12)
23
24
At least 14 bacterial genera are known to contain denitrifying species (Gayle and Benoit 1989;
25
Drysdale et al. 1999). These include Bacillus, Pseudomonas, Methanomonas, Paracoccus, Spirillum,
4-18
Final Draft
1
and Thiobacillus. Denitrification can be accomplished by both heterotrophic and autotrophic organisms
2
(Zumft, 1997). However, most of the denitrifying bacteria are heterotrophic, meaning that they use
3
carbon from organic compounds for cell synthesis as well as energy. There are relatively few species of
4
autotrophic denitrifying bacteria, which obtain carbon for cell synthesis from inorganic compounds. One
5
example is Thiobacillus denitrificans. This organism oxidizes elemental sulfur for energy and obtains
6
carbon for cell biosynthesis from dissolved carbon dioxide or bicarbonate (HCO3–).
7
Denitrification can be accomplished using the carbon from influent organics by creating an anoxic
8
zone or separate anoxic reactor at the head end of the process and recycling nitrified mixed liquor into it.
9
This process is often termed as pre-anoxic denitrification. In post-anoxic denitrification, an external
10
carbon source can be added to the mixed liquor after the ammonia in the wastewater has been oxidized
11
to nitrate or endogenous respiration is used to reduce the nitrates. When very low total nitrogen levels
12
are desired in the final effluent, a combination of pre- and post-anoxic denitrification is often used.
13
External carbon sources include methanol, acetate, ethanol, sugar, butanol, corn syrup, molasses,
14
methane, and industrial wastes such as food processing, breweries, and biodiesel wastes. The advantage
15
of using methanol over other sources in wastewater is that it is free of contaminants such as nitrogen and
16
phosphorus, has the lowest cost and can lead to improved process control and operation. The
17
disadvantage of using methanol intermittently appears to be the initial lag period (several day to weeks)
18
for the growth of methol utilizing denitrifiers (Ginige et al., 2004; Hallin and Pell, 1998; Hallin et al.,
19
1996; Nyberg et al., 1992; Purtschert et al., 1996). The other disadvantage of using methanol is that
20
unlike acetate it cannot be utilized by the phosphorus accumulating organisms (PAOs) in enhanced
21
biological phosphorus removal when both nitrogen and phosphorus removal are desired (deBarbadillo et
22
al. 2008).
23
Simultaneous nitrification and denitrification (SND) can occurs in systems operated at low DO
24
concentrations in the bulk mixed liquor. At the outer periphery of the bioflocs, DO is available for
25
heterotrophs for carbon removal and nitrifiers for nitrification. Inside the floc, however, DO may not
4-19
Final Draft
1
penetrate and anoxic conditions can exist, leading to denitrification. Thus, leading to SND within the
2
aerobic environment. Under these conditions, potentially, neither nitrification nor denitrification may
3
proceed at optimum rates, as low DO can slow down nitrification and the presence of DO can inhibit
4
denitrification.
5
In the SHARON process, NH4+-N is oxidized to NO2–-N in a chemostat reactor (no recycle, SRT =
6
HRT) and the NO2–-N is reduced to nitrogen gas by adding a carbon source under anoxic conditions.
7
The net result of this process is a reduction in theoretical oxygen requirement as NO2–-N is not oxidized
8
to NO3–-N. There is also a reduction in external carbon source requirement because the reduction step
9
from NO3–-N to NO2–-N is eliminated. As discussed earlier, denitrification can also occur in the
10
anammox process (www.anammox.com), where ammonium provides electrons for the denitrification of
11
nitrite to nitrogen gas, as shown below:
12
13
NH4+ + NO2– → N2 + 2H2O
(4.13)
14
15
3.2 Stoichiometry
16
The stoichiometric equations for denitrification depend on the carbon substrate and the source of
17
nitrogen. The energy equations using wastewater and methanol as the carbon substrates and nitrate as
18
the terminal electron acceptor can be written as follows:
19
20
(Wastewater) 10 NO3- + C10H19O3N Æ 5 N2 + 10 CO2 + 3H2O + NH3 + 10 OH-
(4.14)
(Methanol) 6 NO3- + 5 CH3OH Æ 3 N2 + 5 CO2 + 7 H2O + 6 OH-
(4.15)
21
22
23
24
25
The hydroxide ion formed during denitrification reacts with carbon dioxide in the water to create
bicarbonate ions according to the following equation:
4-20
Final Draft
1
2
OH- + CO2 Æ HCO3-
(4.16)
3
4
5
The oxidation-reduction half reactions using oxygen, NO2–-N, and NO3–-N as electron acceptors can
be expressed as follows:
6
7
0.25 O2 + H+ + e– → 0.5 H2O
(4.17)
0.33 NO2– + 1.33 H+ + e– → 0.17 N2 + 0.67 H2O
(4.18)
0.20 NO3– + 1.2 H+ + e– → 0.1 N2 + 0.6 H2O
(4.19)
8
9
10
11
12
13
The significance of the conversion of nitrate to nitrogen gas is that the overall process oxygen demand
14
is reduced by 2.86 g oxygen per g NO3–-N reduced [(0.25×32)/(0.20×14)]. When nitrite is converted to
15
nitrogen gas, oxygen demand is reduced by 1.73 g oxygen per g NO2–-N reduced [(0.25×32)/(0.33×14)].
16
Also, for each equivalent of NO3–-N reduced, one equivalent of alkalinity is produced, which is
17
equivalent to 3.57 g of alkalinity as CaCO3 per g NO3–-N reduced. The stoichiometric relationships for
18
frequently used carbon sources are presented below (Sorensen and Jorgensen, 1993):
19
20
Acetic Acid: 5 CH3COOH + 8 NO3– → 4 N2 +10 CO2 + 6 H2O +8 OH–
(4.20)
Dextrose: 0.208 C6H12O6 + NO3– → 0.5 N2 + 1.25 CO2 + 0.75 H2O
(4.21)
Ethanol: 5 C2H5OH + 12 NO3– → 6 N2 + 10 CO2 + 9 H2O + 12 OH–
(4.22)
21
22
23
24
25
4-21
Final Draft
1
Glycol: 0.50 (CH2OH)2 + NO3– → 0.5 N2 + CO2 + H2O + OH–
(4.23)
Formaldehyde: 1.25 HCHO + NO3– → 0.5 N2 + 1.25 CO2 + 0.75 H2O + OH–
(4.24)
Isoproponol: 0.278 C3H7OH + NO3– → 0.5 N2 + 0.833 CO2 + 0.5 H2O + OH–
(4.25)
2
3
4
5
6
7
Fusel oil (amyl alcohol): 0.167 C5H11OH + NO3– → 0.5 N2 + 0.833 CO2 + 0.5 H2O + OH– (4.26)
8
9
Methane: 8 NO3- + 5 CH4 Æ 4 N2 + 5 CO2 + 6 H2O + 8 OH-
(4.27)
10
11
12
As with nitrification, the inclusion of biosynthesis changes the stoichiometry. The overall result is an
increase in the electron donor (carbon substrate) required per unit mass of nitrate or nitrite reduced.
13
14
3.3 Substrate Requirements
15
A number of factors influence substrate consumption in biological denitrification. The first factor is
16
the concentrations of the electron acceptors present, including nitrate, nitrite, DO, and sulfate (SO42–).
17
Most of the DO present must be reduced before denitrification can proceed. Nitrate and nitrite compete
18
on approximately an equal basis for electrons from the substrate. Sulfate can be reduced biologically, but
19
only after almost all the DO, nitrate, and nitrite have been consumed. Hence, nearly complete
20
denitrification can be obtained without appreciable sulfate reduction.
21
A second factor affecting electron donor requirements is the nature of the donor molecule. Organic
22
compounds are used by bacteria as the source of electrons for energy metabolism, as well as the source
23
of carbon for cell biosynthesis. Inorganic compounds such as molecular hydrogen and sulfur only supply
24
electrons for energy metabolism.
4-22
Final Draft
1
A third factor is the extent of the denitrification reaction. A shortage of electron donor can cause the
2
conversions depicted in Equation 4.15 to stop before nitrogen gas is produced, so that the quantity of
3
NO3–-N removed exceeds the quantity of nitrogen gas produced. The electron donor requirement,
4
expressed in terms of the mass of substrate consumed per unit mass of NO3–-N removed, will then vary
5
directly with the percentage removal of nitrate, up to the point of complete conversion. The total
6
concentration of substrate (as methanol) required to reduce the nitrate, nitrite, and DO present without
7
biosynthesis (Cm) is (McCarty et al., 1969):
8
9
Cm = 2.47 NO3-N + 1.53 NO2-N + 0.87 DO
(4.28)
10
11
Where,
12
Cm = methanol required (mg/L)
13
NO3-N = Initial nitrate nitrogen concentration (mg/L)
14
NO2-N = Initial nitrite nitrogen concentration (mg/L)
15
DO = Dissolved oxygen concentration (mg/L)
16
17
18
Another parameter for evaluating substrate requirements (including the carbon required for growth) is
the substrate consumption ratio (SCR) presented below:
19
20
SCR =
ΔC
ΔNO3eq − N
(4.29)
21
22
Where,
23
ΔC = corresponding change in substrate concentration expressed as (g COD/m3)
24
ΔNO3eq-N = equivalent nitrate concentration consumed (g NO3eq-N/m3)
4-23
(4.30)
Final Draft
1
2
3
The oxygen equivalent of the substrate required (chemical oxygen demand or COD) can be calculated
using the following equation (Tchobanoglous et al., 2003):
4
5
g bsCOD/g NO3–-N = 2.86 / (1–1.42Yobs)
(4.31)
6
7
Where,
8
bsCOD = biodegradable soluble COD (g/d)
9
Yobs = net biomass yield (g biomass VSS produced / g bsCOD removed)
10
11
The subtracte consumption ratio of methanol varies from 3.2 to 6.0 g COD/g NO3eq-N. Substrate
12
consumption ratios reported by Monteith et al. (1980) for industrial organic wastes ranged from 2.2 to
13
10.2 g COD/g NO3eq-N. Typically, a carbon to nitrogen ratio (C/N) of 1.5 to 5 may be required for
14
denitrification to occur effectively, however, a C/N greater than 4 may be required to obtain more than
15
95% nitrate removal in municipal wastewaters treatment systems. Table 4.5 presents a range of C/N
16
optimal for different carbon substrates (Sorensen and Jorgensen, 1993). Table 4.6 presents information
17
on supplemental carbon sources used for denitrification (deBarbadillo et al., 2008).
18
19
20
[INSERT TABLES 4.5 AND 4.6]
21
3.4 Alkalinity Production
22
The quantity of base produced by denitrification can be calculated from the following balanced
23
reaction, as modified from McCarty et al. (1969):
24
25
NO3– + 1.08 CH3OH = 0.065 C5H7O2N + 0.47 N2 + 0.76 CO2 + 1.44 H2O + OH–
(4.32)
26
4-24
Final Draft
1
Therefore, 3.57 mg/L of alkalinity are produced per mg/L NO3–-N reduced when NO3– is used by the
2
denitrifying bacteria for cell synthesis. This alakinity production is beneficial to the overall process to
3
reduce the external addition of alakinity, where required.
4
5
3.5 Kinetics
6
The rate of denitrification has been found to vary depending on the type and concentration of
7
compound used as the carbon source. The availability of soluble readily biodegradable substances results
8
in higher denitrification rates. Denitrification is affected by the DO concentration, pH, temperature, and
9
reactor configuration. Researchers have developed several mathematical models for predicting based on
10
Monod kinetic expression shown below (Henze et al., 1987):
11
12
13
14
15
16
⎛ 1 − YH
rν , NO = ⎜⎜
⎝ 2.86 × YH
Where,
⎛ SS
⎞
⎟⎟ μ mH ⎜⎜
⎠
⎝ KS + SS
⎞⎛
S NO
⎟⎟⎜⎜
⎠⎝ K NO + S NO
⎞⎛
SO
⎟⎟⎜
⎜
⎠⎝ K O , H + S O
⎞
⎟η g X b , H
⎟
⎠
(4.33)
rv,NO = denitrfication rate (g NO3-N reduced/m3.d)
17
μmH = maximum specific growth rate for heterotrophs (d–1)
18
ηg
19
Xb,h = concentration of heterotrophs (mg/L COD)
20
Ss
= concentration of readily degradable organic substrate (mg/L COD)
21
Ks
= half-saturation coefficient for readily degradable substrate (mg/L COD)
22
SO = concentration of dissolved oxygen (mg O2/L)
23
KO,H = half-saturation coefficient for DO in heterotrophic growth (mg O2/L)
24
SNO = concentration of nitrate (mg N/L)
25
KNO = half-saturation coefficient for nitrate (mg N/L)
26
YH = heterotrophic yield (g biomass COD/g substrate COD)
= fraction heterotrophs using nitrate for electron acceptor
4-25
Final Draft
1
2
The values for the biokinetic coefficients for Monod-type expressions are not well defined. Table 4.7
3
summarizes recommended values from two references. The current practice is to use process-simulation
4
models to design denitrfication systems, which in the past was based on empirical rate expressions for a
5
given type of substrate.
6
7
8
9
10
[INSERT TABLE 4.7]
Temperaure has a significant influence on maxium growth rate of denitrifying population, which can
be expressed for methanol utilizing denitrifying bacteria by an Arrhenious equation (Nichols et al., 2007):
11
12
(μmH)T = (μmH)20 (θ)(T-20)
(4.34)
13
14
Where,
15
(μmH)T = maxium specific heterotrophic growth rate at any temperature, T
16
(μmH)20 = maxium specific heterotrophic growth rate at 20oC, and
17
Arrhenous coefficienct, θ = 1.13
18
19
The denitrification rate is strongly affected by the kinetic regime of the reactor. Plug-flow reactors and
20
reactors in series will produce higher denitrification rates when the reaction order is greater than zero.
21
This will typically happen when the availability of substrate limits the denitrification reaction.
22
Denitrifying bacteria grow well under the conditions typically experienced in wastewater, that is, pH
23
between 7 and 8 and temperature between 5 and 25°C. The maximum rate of denitrification is
24
temperature dependent, roughly doubling for every 10°C increase in temperature between 5 and 25°C.
25
Reported values for the temperature coefficient are given in Table 4.8.
26
4-26
Final Draft
1
2
[INSERT TABLES 4.8]
3
Soluble, readily degradable substrates support the highest rates of denitrification. Although methanol
4
is the most typically used soluble substrate, it is not the best on a kinetic basis. Kinetic coefficients for
5
denitrification obtained from a number of reported studies are presented in Tables 4.9 and 4.10. The New
6
York City Department of Environmental Protection evaluated methanol and ethanol addition to improve
7
denitrfication (Fillos et al., 2007). In this study following expressions were determined for specific
8
denitrification rate (SDNR) for acclimated biomass, where ethanol had a higher SDNR value:
9
10
Methanol: (SDNR)T = 0.0738 (1.11)(T-20)
(4.35)
Ethanol: (SDNR)T = 0.161 (1.13)(T-20)
(4.36)
11
12
13
14
Where,
15
SDNR = mg NO3-N removed/mg VSS/day and
16
T = Temperarure in degree Celcius
17
18
Hallin et al. (2006) studied the metabolic profiles and genetic diversity of denitrifying population in
19
the activated sludge by feeding 10 different carbon sources. They observed changes in the community of
20
denitrifiers depending on the carbon source. In addition, a preferrd carbon source could not be identified
21
based on their study since the final outcome depends on conditions specific to an operation, including
22
the costs of supplying carbon source. A different study examined the impact of acetate addition on
23
denitrifying population and found that the activity improved rapidly with the addition of acetate;
24
however, the biomass settleability was adversely affected (Ginige et al., 2005). Therefore, it is important
25
to consider other effects on process operation, when selecting the carbon source for denitrification.
4-27
Final Draft
1
2
3
4
[INSERT TABLES 4.9, AND 4.10]
5
3.6 Toxicity
6
The heterotrophic bacteria that perform denitrification are typically less sensitive to inhibition from
7
toxic chemicals compared to nitrifiers; however, it is still a concern. Oxygen has been found to inhibit
8
nitrite reductase, slowing the rate of nitrite reduction. Oxygen has been found to inhibit nitrite reductase
9
to an even greater extent than nitrate reductase. Hernandez and Rowe (1987) found that nitrite began to
10
accumulate when oxygen was added to a batch denitrifying system and stopped accumulating when the
11
oxygen supply was terminated and the system was flushed with argon gas.
12
Hochstein et al. (1984) conducted experimenst to undertand the influence of oxygen on denitrfication,
13
using a culture Paracoccus halodenitrificans, pure oxygen and a laboratory-scale reactor operating at
14
30oC. They found that, in the absence of DO concentration, the nitrate-limited culture produced nitrogen
15
gas confirming effective denitrification. However, as the oxygen supply was increased, P.
16
halodenitrficans first produced nitrous oxide and then nitrite, indicating the inactivation of nitrous oxide
17
reductase by oxygen and diversion of electrons from nitrite to oxygen, ulmately leading to complete loss
18
of denitrfication. This study also noted that the nitrate reductase was least sensitive to DO, however, it
19
was completely inhibited after 3.3 mg/L of DO concentration in the medium.
20
Excess nitrite concentration can suppress denitrification rates. Rowe et al. (1979) reported that NO2-N
21
concentrations above 14.0 mg/L at a pH of 7.0 inhibited active transport of carbohydrates and amino
22
acid in Pseudomonas aeruginosa, and that concentrations above 350 mg/L completely inhibited active
23
transport by microorganisms in both under anoxic environment with NO3– as the terminal electron
24
acceptor and oxic environment with oxygen as the terminal electron acceptor. Beccari et al. (1979)
25
suggested that inhibition by nitrite is caused by free nitrous acid (HNO2). They found that nitrite reduc-
26
tion rates drop sharply for pH values less than 7.5 in contrast to nitrate reduction rates.
4-28
Final Draft
1
2
3.7 Biofilm Systems
3
Like nitrification, denitrification can also be accomplished in biofilm systems. In fact denitrification
4
is one of the easiest applications for a wide range of biofilm processes, as oxygen transfer will not be
5
limiting factor unlike in the aerobic processes and therefore higher volumertric loadings can be achieved
6
(Rittman and McCarty, 2001). The first filter for denitrification was patented in 1970, suggestiing the
7
application of this technology for a long time (U.S. EPA, 2007b). A list of denitrifying filter
8
manufacturers and the equipment suuplied by them is summarized elsewhere (U.S. EPA, 2007b). Any
9
of the biofilm systems can work efficiently, as long as oxygen transfer is controlled and plugging of the
10
reactor is avoided. The various systems applied for denitrification include (Rittman and McCarty, 2001):
11
•
RBCs in which the air ventilation is controlled
12
•
Submerged fixed beds of rocks, sand, limestone or plastic media
13
•
Fluidized beds of sand, activated carbon, and pellets of ino-exchange resin
14
•
Circulating beds of range of light weight particles
15
•
Membrane reactors where the membrane supplies hydrogen and also the membrane is the
16
support media for biofilm
17
All of the denitrification biofilters are submerged in water. When RBC or packed filter media such as
18
plastic or loose carriers are involved the sludge production is removed via secondary settling. For filter
19
with sand or gravel backwash is emplyed to remove the excess sludge. Moving bed biofilm reactor
20
(MBBR), Biological Anoxic Filter and integrated fixed-film actiavted sludge (IFAS) appear to be some
21
of the widely used reactor configurations for denitrification. The fixed bed IFAS use media such as
22
Ringlace®, Bloweb®, moving bed IFAS use media such as Captor®, Linpor® (sponge), Kaldnes®,
23
Hydroxyl®, Entex (plastic), whereas, the MBBR reactors use media such as Kaldnes®, Entex® or other
24
plastic media (Sen and Randall, 2008). Recently, the Town of Cheshire, Connecticut (13,200 m3/d or
25
3.5 mgd) implemented a, upflow biological anoxic filter with methanol addition to achieve less than 3
4-29
Final Draft
1
mg/L of effluent total nitrogen (Pearson et al., 2008). The MBBR and IFAS technologies were pilot-
2
tested at Noman Cole Jr. Water Pollution Control Plant in Fairfax County, Virginia (Motsch et al.,
3
2007). Both technologies were capable of reducing NOx-N levels from 7 mg/L to under 2 mg/L when
4
operating between 18-20oC. The methanol addition in these trials was at the rate of 5 parts of methanol
5
per part of nitrate-N, which was an over dose maintained intentionally. The nitrate removal rate in the
6
MBBR was equal to 2.5-3.0 g NOx-N/m2/d at 18.5-20oC, resulting in an average effluent NOx-N
7
concentration of 0.6 mg/L. The IFAS system was slightly less effective, achieving effluent NOx-N of
8
0.86 mg/L.
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