ARTICLE IN PRESS
Water Research 37 (2003) 4211–4221
Biological nitrogen removal of high-strength ammonium
industrial wastewater with two-sludge system
J. Carrera*, J.A. Baeza, T. Vicent, J. Lafuente
"
Departament d’Enginyeria Qu!ımica, ETSE, Universitat Autonoma
de Barcelona, Bellaterra 08193, Spain
Received 21 June 2002; received in revised form 28 April 2003; accepted 12 June 2003
Abstract
The biological nitrogen removal (BNR) process is the most common method for removing low quantities of
ammonium from wastewater, but this is not the usual treatment for high-strength ammonium wastewater. The capacity
1
to biologically remove the nitrogen content of a real industrial wastewater with a concentration of 5000 mg N-NH+
4 L
is demonstrated in this work. The experimental system used is based on a two-sludge system, with a nitrifying activated
sludge and a denitrifying activated sludge. This system treated real industrial wastewater for 450 days, and during this
period, it showed the capacity for oxidizing all the ammonium at average nitriﬁcation rates between 0.11 and 0.18 g N1 1
NH+
d . Two key process parameters were evaluated: the maximum nitriﬁcation rate (MNR) and the
4 g VSS
maximum denitriﬁcation rate (MDR). MNR was determined in continuous operation at three different temperatures:
1 1
15 C, 20 C and 25 C, obtaining values of 0.10, 0.21 and 0.37 g N-NH+
d , respectively. Complete
4 g VSS
denitriﬁcation was achieved using two different industrial carbon sources, one containing mainly ethanol and the other
1 1
one methanol. The MDR reached with ethanol (0.64 g N-NO
d ) was about 6 times higher than the MDR
x g VSS
1 1
reached with methanol (0.11 g N-NO
g
VSS
d
).
x
r 2003 Elsevier Ltd. All rights reserved.
Keywords: Biological nitrogen removal (BNR); Nitriﬁcation; Denitriﬁcation; Two-sludge system; Industrial high-strength ammonium
wastewater; External carbon source; Ethanol; Methanol
1. Introduction
There are many different kinds of human activity that
generate wastewater with large quantities of ammonium:
petrochemical, pharmaceutical, fertilizer and food industries, leachates produced by urban solid waste
disposal sites or waste from pig farms. Disposal of this
type of waste is a serious environmental problem
because free ammonia, diluted in water, is one of the
worst contaminators of aquatic life [1].
The BNR process is the most common method for
removing low quantities of ammonium from wastewater, but this is not the usual treatment for high*Corresponding author. Tel.: +34-935812141; fax: +34935812013.
E-mail address: [email protected] (J. Carrera).
strength ammonium wastewater, where physical–chemical systems such as stripping are more frequently used.
The main problem of the biological treatment of highstrength industrial wastewater is that high concentrations of ammonium or nitrite inhibit the nitriﬁcation [2].
However, the BNR process could be an interesting way
for treating high-strength ammonium wastewater from
an environmental and economical point of view [3].
The nitriﬁcation and denitriﬁcation rates are the key
parameters in the design of a biological wastewater
treatment plant with nitrogen removal. For this reason,
it is essential to experimentally determine the maximum
nitriﬁcation rate (MNR) and maximum denitriﬁcation
rate (MDR) in similar conditions, to those of the
treatment plant on an industrial scale [4].
This study aims to deﬁne a biological treatment
system for a real industrial wastewater with an
0043-1354/$ - see front matter r 2003 Elsevier Ltd. All rights reserved.
doi:10.1016/S0043-1354(03)00338-5
ARTICLE IN PRESS
4212
J. Carrera et al. / Water Research 37 (2003) 4211–4221
1
ammonium concentration of 5000 mg N-NH+
4 L . This
concentration is higher than those generally found in the
bibliography (Table 1).
Different conﬁgurations for BNR can be used but, in
order to minimize the total volume of the system, twosludge systems are preferable. This paper presents one of
these systems with two separate stages; the ﬁrst one is a
nitrifying activated sludge stage and the second one is a
denitrifying process. Each stage consists of a reactor and
a settler, which produce the growth of two different
microbial populations: a nitrifying biomass and a
denitrifying biomass.
In order to treat industrial wastewater with a low
COD/N ratio, it is necessary to add an external source of
organic carbon. Different criteria have been used to
choose a speciﬁc external carbon source in the denitriﬁcation process. First, it is necessary to consider
which carbon compound generates the highest MDR.
Published references give contrasting results. Some
authors suggest that acetic acid achieves higher rates
than glucose, methanol or ethanol [9]. However, other
authors achieved similar results with acetic acid to those
achieved with methanol [10,11]. Some references indicate that ethanol reaches higher rates than methanol
[12,13], although another study indicates the opposite
[14].
It is also necessary to consider the costs and
availability of the external carbon source. If the source
is a chemical compound (ethanol, methanol, and acetic
acid), it will be available at a market price. If there are
plans to build an industrial scale treatment plant, the
external carbon source should be cheap and be produced
in sufﬁcient quantities to guarantee the continuous
operation of the wastewater treatment plant. These
requirements can be achieved using carbonaceous byproducts, which are currently generated by an industry,
but that are not considered waste products.
This paper focuses on the study of the nitriﬁcation
and denitriﬁcation stages of a two-sludge system, with
the speciﬁc aim of determining the MNR and the MDR.
In addition, the inﬂuence of temperature on the
nitriﬁcation process and different external carbon
sources on the denitriﬁcation process were also studied.
2. Materials and methods
2.1. Description of wastewaters
In addition to the high-strength ammonium wastewater (N-wastewater), the process included the treatment of a second industrial wastewater that contained
mainly organic matter (COD-wastewater) and useful for
the denitriﬁcation. However, this organic matter was
insufﬁcient to denitrify all the nitrate that had been
Table 1
Industrial high-strength ammonium wastewaters
Wastewater
Concentration
1
(mg N-NH+
4 L )
Reference
Tannery
Sludge dewatering
Leachate
Landﬁll leachate
200–500
600–700
1200
2200
[5]
[6]
[7]
[8]
Table 2
Basic composition of both real industrial wastewaters
Component
N-wastewater
(mg L1)
COD-wastewater
(mg L1)
COD
N-NH+
4
F
Cl
SO2
4
0
4000–6000
30–50
500–600
15000–20000
1300–1500
0
0
700–1000
300–800
formed. Therefore, an external carbon source was also
used.
The basic composition of the two real industrial
wastewaters are summarized in Table 2. As can be seen,
the concentration of ammonium in the N-wastewater
1
ranged from 4000 to 6000 mg N-NH+
4 L , while the
concentration of COD was 1300–1500 mg COD L1 in
the COD-wastewater. Most of this organic matter was
ethanol, and thus easily biodegradable. The N-wastewater also contained high concentrations of chloride
and sulphate anions.
2.2. Description of the external carbon sources
Two external carbon sources were used, both
by-products of two industrial processes. The ﬁrst
by-product was a mixture of waste from alcoholic
drinks production, and the main carbon source was
ethanol. The second by-product was a waste from a
chemical industry, made up of a mixture of methanol,
isopropylic alcohol and acetone. Its main component
was methanol. Table 3 shows the composition of the two
by-products. Throughout this study, they have been
called the ‘ethanol mixture’ (that formed by waste from
alcoholic drinks) and the ‘methanol mixture’ (that
formed from chemical waste).
2.3. Monitoring and control system
Every reactor of the treatment plant had in-line
sensors (dissolved oxygen (DO), pH, ORP, temperature)
connected to probe controllers. All these controllers and
the mechanical elements of the pilot plant were
ARTICLE IN PRESS
J. Carrera et al. / Water Research 37 (2003) 4211–4221
itoring, data backup, and control of key process
parameters (ﬂows, pH, DO and temperature). The pH
control was based on an ON/OFF-algorithm acting over
a solid dispenser that adds sodium carbonate. The DO
control was based on a digital PID algorithm programmed in the computer, which modiﬁes the airﬂow
using a mass ﬂowmeter (Bronkhorst Hi-Tec, 0–20
Ln min1). The treatment plant is located indoors,
permitting temperature regulation through a heating,
ventilation and air-conditioning system.
connected to a PC, through different data acquisition
cards (Advantech PCL726, PCL813 and PCLD885).
Speciﬁc software was developed in C language in order
to automate all the system. It was based on previous
software developments [15] and included graphic mon-
Table 3
Percentage composition of the by-products used as external
carbon sources
By-product
Component
Waste from alcoholic drinks
(‘ethanol mixture’)
Ethanol
8.7
Glucose
Glycerol
Water
0.8
0.5
90.0
Methanol
60.0
Acetone
Isopropylic
alcohol
Water
10.0
10.0
Chemical waste (‘methanol
mixture’)
4213
Percentage
(in weight)
2.4. Experimental settings
The experimental work was carried out in a twosludge, treatment plant with two separate stages:
nitriﬁcation and denitriﬁcation. Fig. 1 shows a diagram
of the treatment plant. The N-wastewater was fed into
the nitrifying activated sludge system, made up of a 27 L
aerobic reactor and a settler. The aerobic reactor used
the automatic pH control with a setpoint of 7.5. The DO
was maintained at 3 mg O2 L1 through the PID
controller. The temperature was maintained at 20 C
for 300 days, then it was changed at 15 C for 30 days
20.0
Alkalinity recirculation
Carbonate
dosage
ORP
ORP
DO
pH
DO
pH
Temp
Temp
Effluent
Settler
Settler
N-wastewater
Aerobic
Anoxic
Nitrogen gas
stripping
Sludge recirculation
Sludge recirculation
Stirrers
Mass
flowmeter
COD-wastewater
Frequency
converters
External carbon source
Probes
Air
Pumps
PC
Water and sludge line
Air line
Signal line
Fig. 1. Layout of the treatment plant.
pH control
ARTICLE IN PRESS
4214
J. Carrera et al. / Water Research 37 (2003) 4211–4221
and ﬁnally changed to 25 C. The sludge retention time
(SRT) was kept at around 25 days.
The efﬂuent of the nitrifying activated sludge system
was one of the three inﬂuents to the denitrifying
activated sludge system (Fig. 1). The other two were
the COD-wastewater and the external carbon source.
The denitriﬁcation stage was made up of a 27 L reactor
without aeration; a 15 L aerated tank and a settler. The
nitrogen gas formed in the anoxic reactor was stripped
in the aerated tank, thus favouring the sedimentation
stage that follows. The temperature was the same as for
the aerobic reactor. The pH was not controlled and its
value was about 8.0–9.0. The SRT was kept at around
15 days.
The nitrogen-loading rate (NLR), nitriﬁcation and
denitriﬁcation rates were deﬁned as
N NHþ
4 in
NLR ¼
;
ð1Þ
HRTreactor ½VSSreactor
þ
N NHþ
4 in N NH4 out
;
ð2Þ
rnitrification ¼
HRTreactor ½VSSreactor
N NO
x in N NOx out
;
ð3Þ
rdenitrification ¼
HRTreactor ½VSSreactor
where HRT is the hydraulic retention time, [VSS]reactor
the biomass concentration, [N-NH+
4 ]in the inﬂuent
ammonium concentration, [N-NH+
4 ]out the efﬂuent
ammonium concentration, [N-NO
x ]in the inﬂuent oxidized nitrogen concentration and [N-NO
x ]out the
efﬂuent oxidized nitrogen concentration.
2.5. Analytical methods
The analysis of total suspended solids (TSS), volatile
suspended solids (VSS), sludge volumetric index (SVI),
alkalinity and ammonium were carried out using the
methodology described in APHA’s Standard Methods
[16]. The analyses of chloride (Cl), sulphate (SO2
4 )
nitrite (NO
2 ), nitrate (NO3 ) and ﬂuoride (F ) were
carried out by capillary electrophoresis, using a
WATERS Quanta 4000E CE. The electrolyte used was
a WATERS commercial solution. The conditions of the
analysis were temperature of 20 C, 15 kV from a
negative source, indirect UV detection at 254 nm and
5 min of analysis.
3. Results and discussion
The nitrifying and denitrifying reactors of the present
study were inoculated with biomass emanating from a
preliminary study using a modified Ludzack-Ettinger
conﬁguration that treated these industrial wastewaters
for a year [17]. Both systems started operating with the
same microbial population and then developed towards
specialized biomass (owing to the speciﬁc experimental
conditions). The two-sludge plant was operated continuously for 450 days, using real, high-strength
industrial wastewater.
3.1. Nitrification
3.1.1. pH control
The N-wastewater contained less alkalinity than
stoichiometrically required for nitriﬁcation. In order to
solve this problem, an automatic system for pH control
was installed for adding solid sodium carbonate to the
nitrifying reactor. This control supplies the necessary
alkalinity so that the pH conditions favour the process.
The optimal pH for the nitriﬁcation process is in the
range of 7.5–8.5 [18]. It was chosen a value of 7.5, since
at higher pH, the equilibrium ammonium–ammonia
would be displaced to ammonia, which can inhibit the
process. The addition of sodium carbonate also provided suitable conditions for the nitrifying microorganisms, by avoiding potential restrictions in the use of one
of their substrates: inorganic carbon.
The solid dispenser was calibrated to give an exact
amount of sodium carbonate (1.3 g Na2CO3 per dosage).
As the control system kept a register of the total dosages
per day, it was possible to quantify the total amount of
carbonate used per day. During the experiments, it was
checked to see if the carbonate consumption agreed with
the theoretical values of alkalinity destruction, acceptable for design of nitriﬁcation systems (7.1 g CaCO3 per
2
+
g N-NH+
4 consumed, or 4.26 g CO3 /g N-NH4 ) [18].
The experimental value obtained was 6.170.6 g Na2CO3
per g N-NH+
consumed, which corresponds to
4
+
4.470.4 g CO2
3 /g N-NH4 (a value which is very close
to the theoretical).
From day 140 of operation, part of the water coming
out of the denitrifying system was recirculated back to
the nitrifying system. Part of the alkalinity generated by
the denitrifying system is recovered with this recirculation, thus decreasing the amount of added carbonate.
3.1.2. Sludge evolution
The initial biomass concentration of the nitrifying
reactor was about 4000 mg VSS L1. Fig. 2 shows the
VSS and TSS concentrations of this reactor and its SVI.
The VSS concentration was kept around 35007700 mg
VSS L1 throughout the study. The VSS/TSS ratio was
around 4878%. Nevertheless, this ratio decreased
below the average value between 200 and 250 days; this
decrease was due to the change in the external alkalinity
source. The original sodium carbonate was changed to
another cheaper source, calcium hydroxide; the addition
of calcium ion in the nitrifying reactor caused the
precipitation of sulphate and chloride salts. These salts
increased the inorganic composition of the sludge and
ARTICLE IN PRESS
J. Carrera et al. / Water Research 37 (2003) 4211–4221
100
14000
TSS
VSS
% VSS/ TSS
average % VSS/ TSS
SVI
80
70
60
8000
50
6000
40
30
4000
150
100
SVI (mL·g-1)
10000
200
90
VSS/ TSS percentage
12000
-1
TSS and VSS (mg·L )
4215
50
20
2000
10
0
0
0
50
100
150
200
250
300
350
400
0
450
Time (days)
Fig. 2. TSS, VSS, % VSS/TSS and SVI of the nitrifying sludge system throughout the study.
7000
without recirculation
with recirculation
90
5000
[N] (mg·L-1)
80
[N-NH4+ ] influent
[N-NH4+ ] effluent
[N-NOx- ] effluent
4000
70
60
% ammonium removal
50
3000
40
2000
30
20
1000
10
0
0
50
100
150
200
250
300
350
400
Percentage of ammonium removal
100
6000
0
450
Time (days)
Fig. 3. Nitrogen concentration in the inﬂuent and the efﬂuent of the nitrifying sludge system and the percentage of ammonium
removal throughout the study.
the VSS/TSS ratio decreased by 30–35%. In order to
correct this problem, the alkalinity source was changed
again to sodium carbonate and the VSS/TSS ratio
reached the average value of the study. The nitrifying
sludge showed good settleability throughout the study,
with SVI values below 50 mL g1.
3.1.3. Nitrogen removal
Fig. 3 shows the concentrations of all nitrogen
compounds in the inﬂuent and the efﬂuent of the
nitrifying system and the ammonium removal percentage during the 450 days of operation. There are two
periods, the ﬁrst one without alkalinity recirculation and
the second one with alkalinity recirculation (the efﬂuent
of the denitrifying system, see Fig. 1). This recycle
involved the dilution of the inﬂuent ammonium con-
centration of the nitrifying system, but not a decrease in
the NLR. Table 4 shows the operational parameters of
the two runs. The average nitriﬁcation rate was very
high in both runs. The ammonium removal percentage
was also very high throughout the study, ranging from
90% to 100%.
The comparison of ammonium concentration in the
inﬂuent with oxidized nitrogen concentration (nitrate
plus nitrite) in the efﬂuent, conﬁrmed that the ammonium removal is due to the nitriﬁcation process,
since both concentrations are about equal. In
addition, the carbonate consumption obtained (Section
3.1.1) agreed with the nitriﬁcation stoichiometry. These
results reject the concept of ammonia removal by
stripping or other biological ways like ANAMMOX
process [19].
ARTICLE IN PRESS
J. Carrera et al. / Water Research 37 (2003) 4211–4221
4216
Table 4
Operational parameters for the nitrifying system throughout the study
Run
Period
(days)
[N-NH+
4 ]inﬂuent
(mg N L1)
[N-NH+
4 ]efﬂuent
(mg N L1)
Inﬂuent ﬂow
(L d1)
HRT
(days)
VSS
(mg L1)
Average nitriﬁcation rate
1 1
(g N-NH+
d )
4 g VSS
1
2
140
310
48007450
19007400
80750
40720
2.7–3.0
5.4–9.0
9–10
3–5
35007500
35007800
0.13–0.15
0.11–0.18
1000
Nitrification rate
NLR
[N-NH4+ ] influent
0.05
500
[N-NH4+ ] effluent
0.00
0
0
5
(a)
10
Time (days)
15
20
2500
2000
0.3
1500
0.2
1000
Nitrification rate
NLR
[N-NH4+ ] influent
0.1
500
[N-NH4+ ] effluent
0.0
0
0
5
(b)
10
Time (days)
15
20
0.6
2000
0.5
1500
0.4
0.3
1000
Nitrification rate
NLR
[N-NH4+ ] influent
0.2
500
[N-NH4+ ] effluent
0.1
0.0
0
0
(c)
[N-NH4+ ] (mg·L-1)
NLR and nitrification rate
-1 -1
(g N-NH4+·g VSS ·d )
0.4
[N-NH4+ ] (mg·L-1)
NLR and nitrification rate
-1 -1
(g N-NH4+·g VSS ·d )
1500
0.10
+
-1
[N-NH4 ] (mg·L )
2000
0.15
NLR and nitrification rate
-1 -1
(g N-NH4+·g VSS ·d )
3.1.4. Maximum nitrification rate
The experiments to determine the MNR were done
under similar conditions to those of an industrial scale
treatment plant, i.e. continuous. The problem with
doing these experiments was the high nitrogen concentration in the N-wastewater, because if the NLR is
higher than the MNR, a large accumulation of
ammonium can take place. This accumulation can
inhibit the process [2] and the observed nitriﬁcation rate
would not be the maximum. Therefore, the experiments
were carried out with a gradual and controlled increase
in the NLR, so that the nitriﬁcation rate was the as close
as possible to the NLR. Only when the NLR was slightly
above the MNR, was there some accumulation of
ammonium in the system; however, the observed
nitriﬁcation rate was still at maximum.
Three experiments were carried out, at 15 C, 20 C
and 25 C. Fig. 4(a) shows the results of the ﬁrst
experiment in which the temperature was 15 C. The
experiment began with a NLR of 0.06 g N-NH+
4 g
VSS1 d1. Since no accumulation of ammonium
occurred, the NLR was increased to 0.13 g N-NH+
4 g
VSS1 d1. This NLR was clearly higher than the MNR
1
at 15 C, since accumulation of 150 mg N-NH+
4 L
was produced. The MNR was calculated as an average
of the rates measured between days 5 and 16, i.e. a
period of three HRTs, and was found to be 0.1070.01 g
1 1
N-NH+
d . The results of the second and the
4 g VSS
third experiments are shown in Figs. 4(b) and (c),
respectively. The MNR in both experiments was
evaluated using the same procedure of the ﬁrst experiment. The MNR in the second experiment (T ¼ 20 C)
was calculated using the rates measured between 13 and
20 days, i.e. a period of two HRTs, and was found to be
1 1
0.2170.01 g N-NH+
d . The MNR in the
4 g VSS
third experiment (T ¼ 25 C) was calculated using the
rates obtained between 12 and 20 days, i.e. a period of
two HRTs, and was found to be 0.3770.03 g N-NH+
4 g
VSS1 d1. These values show the inﬂuence of temperature on nitriﬁcation rates. The temperature coefﬁcient
obtained, adjusting these rates to an Arrenhius-type
equation (rnitT1 ¼ rnitT2 yðT1 T2 Þ ), is y ¼ 1:1470:03:
Table 5 compares the MNR obtained at 25 C in this
study, with different published dates of systems for
treating high-strength, ammonium wastewaters. The
MNR at 25 C was chosen since most published data
5
10
Time (days)
15
20
Fig. 4. Experimental determination of the maximum nitriﬁcation rate at: (a) 15 C, (b) 20 C and (c) 25 C.
was obtained at that temperature. The table shows that
this study’s MNR is clearly higher than those for BNR
single-sludge systems. This is a logical result, since in the
single-sludge systems there is a negative inﬂuence of the
inﬂuent COD/N ratio on the achievable MNR [17,25].
ARTICLE IN PRESS
J. Carrera et al. / Water Research 37 (2003) 4211–4221
4217
Table 5
Nitriﬁcation rates in systems for treating high-strength ammonium wastewater
Technology
Nitrifying bioﬁlm
Nitrifying bioﬁlm
Nitrifying bioﬁlm
Nitrifying bioﬁlm
Nitrifying sludge
Nitrifying sludge
Single-sludge
Single-sludge
Single-sludge
T ( C)
30
30
28
28–32
28–32
25
27
25
25
Nitriﬁcation rate
Reference
1 1
(g N-NH+
d )
4 g VSS
3 1
(g N-NH+
d )
4 m
0.35
—
—
0.18–0.21
0.14–0.18
0.37
0.16
0.05
0.03
—
5.0
0.55
0.8–1.0
0.6–0.8
1.3
—
—
—
More speciﬁcally, the MNR obtained with the
1 1
nitrifying system, 0.37 g N-NH+
d , is 12 times
4 g VSS
higher than the MNR obtained with the same wastewater in a single-sludge system, with nitriﬁcation–
1 1
denitriﬁcation (0.03 g N–NH+
d ) with an
4 g VSS
inﬂuent COD/N ratio of 3.4 [17]. In order to compare
the MNR of this study with the nitrifying bioﬁlm rates, a
nitriﬁcation rate is needed by unit of volume and not by
mass unit. In the experiment at 25 C, the MNR by unit
of volume was 1.3 g N m3 d1. Comparison of data
from this work to that of immobilized nitrifying biomass
systems, provides an even greater disparity in the results.
The reason for such different results might be the
heterogeneity of the industrial wastewater. Although
they are all classiﬁed as high-strength ammonium
wastewaters, each one has its own peculiarities that
inﬂuence a biological treatment process. The industrial
wastewater studied here contains a high concentration
of ammonium, as well as high concentrations of
sulphate, chlorine and a certain amount of ﬂuoride
(Table 2). These components can inﬂuence biological
treatment; for example, high chlorine [5] or ﬂuoride [26]
concentrations can inhibit the nitriﬁcation process.
3.1.5. Inhibition of the nitrification by substrate
The results obtained during the ﬁrst run conﬁrmed the
difﬁculty of working with an inﬂuent concentration of
1
5000 mg N-NH+
4 L , since a small drop in the removal
percentage resulted in an accumulation of up to 500 mg
1
N-NH+
in the nitrifying reactor. The reason for the
4 L
decrease in the removal percentage was a direct result of
the NLR being increased above the nitriﬁcation rate of
the system. The free ammonia concentration, at 500 mg
1
N-NH+
4 L , 20 C and pH=7.5, was about 7.7 mg
NH3 L1. This free ammonia concentration can cause
the inhibition of the ammonium oxidizing and the nitrite
oxidizing bacteria [8]. This inhibition also resulted in the
1
accumulation of 1500 mg N-NO
and the decrease
2L
in the nitriﬁcation rate, from 0.15 to 0.10 g N1 1
NH+
d . In order to mitigate this inhibition,
4 g VSS
[20]
[21]
[22]
[6]
[6]
This study
[23]
[24]
[17]
the NLR was decreased when the ammonium concentration in the nitrifying reactor reached values ap1
proaching 300 mg N-NH+
4 L .
3.2. Denitrification
The efﬂuent of the nitriﬁcation system was treated in
the denitrifying system, using two different external
carbon sources, during different periods. In order to
study the effects of both external carbon sources on the
denitriﬁcation rate, the MDR of the system, in steady
state, was evaluated.
3.2.1. Denitrification with ‘ethanol mixture’ as external
carbon source
The ﬁrst external carbon source to be tested was the
‘ethanol mixture’. This external carbon source contained
mainly the same organic matter as the COD-wastewater
that the industry generates. Fig. 5(a) shows the inﬂuent
and efﬂuent nitrogen concentrations (nitrate plus nitrite,
although nitrite contribution was almost negligible).
The temperature was maintained at 20 C and the
external carbon source ﬂow was adjusted to produce an
inﬂuent COD/N ratio of 5 g COD g N1, thus ensuring
that the system was not limited by the organic matter.
The 200 days of the external carbon source study were
divided into six runs. Table 6 summarizes the values of
the operational parameters and gives the average value
of the nitrogen and biomass concentrations for each run,
along with the standard deviation. Errors in the
denitriﬁcation rate were calculated on the assumption
that the error associated with each concentration was
the standard deviation.
During runs 1 and 2, the system was limited by the
amount of nitrogen supplied by the nitriﬁcation system.
In order to operate at faster rates, solid sodium nitrate
was added quantitatively to the denitrifying reactor in
runs 3–6.
In runs 3 and 4, the system was still limited by the
substrate since the efﬂuent nitrogen concentration
ARTICLE IN PRESS
J. Carrera et al. / Water Research 37 (2003) 4211–4221
4218
1
ﬂuctuated between 0 and 30 mg N-NO
x L . During run
4, the biomass concentration in the reactor increased to
an average value of 9700 mg VSS L1. This high biomass
concentration produced a large amount of nitrogen gas,
which provoked rising problems in the settler. In run 5,
this problem was overcome by a reduction in the
biomass concentration in the system, down to an
average value of 3600 mg VSS L1. In this run, the
denitriﬁcation was not limited by substrate, since the
efﬂuent nitrogen concentration always exceeded 60 mg
1
N-NO
x L . The denitriﬁcation rate obtained under
3000
[N-NOx- ] (mg·L-1)
5
[N-NOx- ]influent
[N-NOx- ] effluent
2500
Average concentration
2000
4
1500
6
3
1000
2
1
500
0
0
50
100
Time (days)
(a)
150
200
[N-NOx- ] (mg·L-1)
2500
[N-NOx- ]influent
[N-NOx- ] effluent
2000
Average concentration
3
1500
2
1
1000
500
0
0
(b)
20
40
80
60
Time (days)
100
these conditions was the MDR: 0.6470.10 g N-NO
x g
VSS1 d1. This rate was maintained for 21 days, so it
can be assumed that the steady state was reached.
Operating at such a high rate led to operational
problems due to the large quantities of nitrogen gas
produced. For this reason, the inﬂuent concentration
was reduced. In these new conditions the denitriﬁcation
1 1
rate dropped to 0.2270.07 g N-NO
d and
x g VSS
stayed at that value for 55 days with a high performance
level, without causing any major operational difﬁculties.
This last run showed that this system can operate at high
denitriﬁcation rates for an extended period. The COD/N
ratio consumed during denitriﬁcation, using the ‘ethanol
mixture’, was 4.370.4 g COD g N1.
Table 7 compares the MDR obtained by this study
with those of several published works. It can be seen that
the MDR of this study were higher than those obtained
by single-sludge systems, with nitriﬁcation–denitriﬁcation. The reason is that in this system, no oxygen is sent
to the anoxic reactor so there is no aerobic consumption
of organic matter. This produces a higher percentage of
denitrifying bacteria in the system. However, the MDR
reached in batches, with pure cultures of denitrifying
bacteria, is higher than that of this study.
120
140
Fig. 5. Inﬂuent and efﬂuent nitrogen concentrations of the
denitrifying system. (a) With ‘ethanol mixture’ as a carbon
source and (b) with ‘methanol mixture’ as carbon source.
3.2.2. Denitrification with ‘methanol mixture’ as external
carbon source
From day 200 of operation, the external carbon
source was changed to the ‘methanol mixture’. It was
diluted with tap water in order to maintain the same
HRTs employed with the ‘ethanol mixture’. Fig. 5(b)
shows inﬂuent and efﬂuent nitrogen concentrations
(nitrate plus nitrite, although nitrite contribution was
almost negligible) of the denitrifying system with
‘methanol mixture’ as the carbon source.
The temperature was kept at 25 C and the inﬂuent
COD/N ratio at 5 g COD g N1, to ensure that the
system was not limited by the organic matter. The 140
days of operation with this external carbon source were
divided into three periods. Table 8 summarizes the
values of the operational parameters and shows the
average values of the nitrogen and biomass concentrations
Table 6
Operational parameters for the denitrifying system with ‘ethanol mixture’
Run
Period
(days)
[N-NO
x ]inﬂuent
(mg N L1)
[N-NO
x ]efﬂuent
(mg N L1)
Inﬂuent ﬂow
(L d1)
HRT
(days)
VSS
(mg L1)
Average denitriﬁcation rate
1 1
(g N-NO
d )
x g VSS
1
2
3
4
5
6
50
41
16
13
21
55
450750
750760
1240730
1820720
2630740
16007120
46710
40720
33720
17710
110760
30710
19.1
24.7
26.3
35.0
38.2
28.0
2.2
1.7
1.6
1.2
1.1
1.5
80007300
84007400
81007500
97007600
36007200
47007300
0.0270.02
0.0570.02
0.0970.03
0.1570.03
0.6470.10
0.2270.07
ARTICLE IN PRESS
J. Carrera et al. / Water Research 37 (2003) 4211–4221
4219
Table 7
Denitriﬁcation rates with ethanol
Technology
Process
T ( C)
Denitriﬁcation rate
1 1
(g N-NO
d )
x g VSS
Reference
Single-sludge
Single-sludge
Single-sludge
Denitrifying sludge
Denitrifying pure culture
Continuous
Continuous
Continuous
Continuous
Batch
20
20
20
20
25
0.02–0.12
0.29a
0.39a
0.64
3.3
[14]
[27]
[28]
This study
[12]
a
Original experimental data at 15 C, corrected using a temperature coefﬁcient of 1.10.
Table 8
Operational parameters for the denitrifying system with ‘methanol mixture’
Run
Period
(days)
[N-NO
x ]inﬂuent
(mg N L1)
[N-NO
x ]efﬂuent
(mg N L1)
Inﬂuent ﬂow
(L d1)
HRT
(days)
VSS
(mg L1)
Average denitriﬁcation rate
1 1
(g N-NO
d )
x g VSS
1
2
3
19
64
57
900730
10907100
13507200
590730
5207200
90760
23.3
21.0
32.2
1.8
2.0
1.3
47007100
35007200
58007300
0.0470.03
0.0870.04
0.1770.06
Table 9
Denitriﬁcation rates with methanol
Technology
Process
T ( C)
Denitriﬁcation rate
1 1
(g N-NO
d )
x g VSS
Reference
Single-sludge
Single-sludge
Single-sludge
Single-sludge
Denitrifying sludge
Denitrifying sludge
Denitrifying pure culture
Continuous
Continuous
Continuous
Continuous
Continuous
Batch
Continuous
Batch
Batch
30
10
20
—
25
22
20
25
25
0.14
0.06
0.17–0.48
0.03
0.17
0.29
0.55
1.3
2.2
[3]
[10]
[14]
[29]
This study
[11]
[30]
Denitrifying pure culture
in each of the periods, along with the standard
deviation. Errors in the denitriﬁcation rate were
calculated assuming that the error associated to each
concentration is the standard deviation. With this
carbon source, there was no sign of limitation by
nitrogen, so it was not necessary to add solid sodium
nitrate to achieve the MDR.
The average denitriﬁcation rate increased from
1 1
0.0470.03 g N-NO
d
to 0.1770.06 g N–
x g VSS
1 1
NO
g
VSS
d
in
about
80 days. This increase was
x
due to the acclimation time needed for the denitrifying
biomass to the ‘methanol mixture’. This period agrees
with the acclimation time to methanol, which has been
reported to be between 50 and 100 days [10,11]. The
COD/N ratio consumed, using the ‘methanol mixture’,
was 3.970.5 g COD g N1.
Table 9 compares the MDR obtained using the
‘methanol mixture’ with different published studies that
used pure methanol. The range of denitriﬁcation rates in
[12]
single-sludge systems with nitriﬁcation–denitriﬁcation is
quite large, with values higher and lower than those of
this study. The results for denitrifying sludge systems
and pure cultures are all higher than those of the current
study. This lower rate can be due to the carbon source
used for the present study, since pure methanol was not
utilized.
3.2.3. Comparison of maximum denitrification rates
In order to compare the MDR obtained with the two
carbon sources, it was necessary to correct the MDR of
the ‘methanol mixture’, since that rate was determined at
25 C and that of the ‘ethanol mixture’ at 20 C. The
temperature coefﬁcient, determined experimentally with
the same biomass [31], was 1.10. Using that correction,
the MDR at 20 C for the ‘methanol mixture’ would be
1 1
about 0.11 g N-NO
d , while for the ‘ethanol
x g VSS
1 1
mixture’, it was 0.64 g N-NO
d . Therefore,
x g VSS
the MDR with ‘ethanol mixture’ appears to be about 6
ARTICLE IN PRESS
4220
J. Carrera et al. / Water Research 37 (2003) 4211–4221
times higher than the MDR with the ‘methanol mixture’.
This ratio is higher than other bibliographic values
[12,13], in which MDR with ethanol is about 1.5–3.5
times higher than MDR with methanol. This difference
can be due to the composition of the ‘methanol mixture’
used, which includes acetone (10%) and isopropylic
alcohol (10%).
4. Conclusions
The results of this study show that it is possible to
remove the ammonium of an industrial wastewater with
1
up to 5000 mg N-NH+
and high concentration of
4 L
sulphates and chlorides, using a biological, two-sludge
system with nitriﬁcation and denitriﬁcation.
The nitriﬁcation rate achieved with the activated
sludge nitrifying system was very high. The MNR
obtained at 15 C, 20 C and 25 C were 0.10, 0.21 and
1 1
0.37 g N-NH+
d , respectively. The nitrifying
4 g VSS
sludge showed good settleability throughout the study,
with SVI values below 50 mL g1. The ammonium
removal percentage was very high throughout (90–
100%), regardless of the concentration of the inﬂuent.
The main problem for the treatment of this industrial
wastewater, with a high concentration of ammonium,
was possible inhibition by substrate of the nitriﬁcation
process.
The efﬂuent of the nitrifying system was treated in the
activated sludge denitrifying system, using two different
external carbon sources, during different periods. The
ﬁrst carbon source was mainly ethanol and the second
one was a mixture of methanol (60%), acetone (10%)
and isopropylic alcohol (10%). In order to study the
effects of both external carbon sources on the denitriﬁcation rate, the MDR in steady state was evaluated.
The MDR obtained with the ﬁrst carbon source was
1 1
about 0.64 g N-NO
d at 20 C, while the
x g VSS
MDR obtained with the second carbon source was
1 1
0.17 g N-NO
d at 25 C. If the temperature
x g VSS
coefﬁcient is taken into account, the denitriﬁcation rate
with ethanol was about six times higher than that
obtained with methanol.
Acknowledgements
This work has been supported by CICYT (project
REN2000-0670/TECNO). The Generalitat de Catalunya
provided ﬁnancial support for Julia! n Carrera through a
pre-doctoral fellowship. We wish to express our
gratitude to the company FREIXENET S.A. The
Departament d’Enginyeria Qu!ımica (UAB) is a member
of the CERBA.
References
[1] Efﬂer SW, Brooks CM, Auer MT, Doerr SM. Free
ammonia and toxicity criteria in a polluted urban lake.
J Water Pollut Control Fed 1990;62:71–9.
[2] Anthonisen AC, Loehr RC, Prakasan TBS, Srineth EG.
Inhibition of nitriﬁcation by ammonia and nitrous acid.
J Water Pollut Control Fed 1976;48(5):835–52.
[3] Teichgraber B, Stein A. Nitrogen elimination from sludge
treatment reject water comparison of the steam-stripping
and denitriﬁcation processes. Water Sci Technol
1994;30(6):41–51.
[4] Harrem.oes P, Haarbo A, Winther-Nielsen M, Thirsing C.
Six years of pilot plant studies for design of treatment
plants for nutrient removal. Water Sci Technol
1998;38(1):219–26.
.
[5] Orhon D, Genceli EA, Sozen
S. Experimental evaluation
of the nitriﬁcation kinetics for tannery wastewaters. Water
SA 2000;26:43–50.
.
[6] Arnold E, Bohm
B, Wilderer PA. Application of activated
sludge and bioﬁlm sequencing batch reactor technology to
treat reject water from sludge dewatering systems: a
comparison. Water Sci Technol 2000;41(1):115–22.
[7] Shiskowski DM, Mavinic DS. Biological treatment of a
high ammonia leachate: inﬂuence of external carbon
during initial start-up. Water Res 1998;32(8):2533–41.
[8] Ilies P, Mavinic DS. The effect of decreased ambient
temperature on the biological nitriﬁcation and denitriﬁcation of a high ammonia landﬁll leachate. Water Res
2001;35:2065–72.
[9] Constantin H, Fick M. Inﬂuence of C-sources on the
denitriﬁcation rate of high-nitrate concentrated industrial
wastewater. Water Res 1997;31(3):583–9.
[10] Nyberg U, Aspergren H, Andersson B, Jansen J la C,
Villadsen IS. Full-scale application of nitrogen removal
with methanol as carbon source. Water Sci Technol
1992;26(5-6):1077–86.
[11] Lee S, Koopman B, Park S, Cadee K. Effect of fermented
wastes on denitriﬁcation in activated sludge. Water
Environ Res 1995;67(7):1119–22.
[12] Christensson M, Lie E, Welander T. A comparison
between ethanol and methanol as carbon sources for
denitriﬁcation. Water Sci Technol 1994;30(6):83–90.
[13] Andersson B, Aspergren H, Nyberg U, la Cour Jansen J,
Odegaard H. Increasing the capacity of an extended
nutrient removal plant by using different techniques.
Water Sci Technol 1998;37(9):175–83.
[14] Henze M. Capabilities of biological nitrogen removal
processes from wastewater. Water Sci Technol
1991;23:669–79.
[15] Baeza JA, Gabriel D, Lafuente J. Improving the nitrogen
removal efﬁciency of an A2/O based WWTP by using an
on-line Knowledge Based Expert System. Water Res
2002;36(8):2109–23.
[16] APHA. Standard methods for the examination of water
and wastewater, 19th ed. Washington, DC: American
Publishers Health Association, 1995.
[17] Carrera J, Vincent T, Lafuente J. Effect of inﬂuent COD/N
ratio on biological nitrogen removal (BNR) from highstrength ammonium industrial wastewater. Process Biochem 2003; in press.
ARTICLE IN PRESS
J. Carrera et al. / Water Research 37 (2003) 4211–4221
[18] EPA Manual of Nitrogen Control. EPA/625/R-93/010 US
Environmental Protection Agency, Washington, USA,
1993.
[19] Jetten M, Strous M, van Dogen U, van de Pas-Schoonen
K, Schalk J, van Dongen U, Logemann S, Muyzer G, van
Loosdrecht M, Kuenen G. The anaerobic oxidation of
ammonium. FEMS Microbiol Rev 1999;22:421–37.
[20] van Benthum WAJ, Derissen BP, van Loosdrecht MCM,
Heijnen JJ. Nitrogen removal using nitrifying bioﬁlm
growth and denitrifying suspended growth in a bioﬁlm
airlift suspension reactor coupled with a chemostat. Water
Res 1998;32(7):2009–18.
[21] Tijhuis L, Huisman JL, Hekkelman HD, Van Loosdrecht
MCM, Heijnen JJ. Formation of nitrifying bioﬁlms on
small suspended particles in airlift reactors. Biotechnol
Bioeng 1995;47:585–95.
[22] Fdz-Polanco F, Villaverde S, Garc!ıa PA. Temperature
effect on nitrifying bacteria activity in bioﬁlters: activation
and free ammonia inhibition. Water Sci Technol
1994;30(11):121–30.
[23] Gupta SK, Sharma R. Biological oxidation of high strength
nitrogenous wastewater. Water Res 1996;30(3):593–600.
[24] Battistoni P, Morini C, Pavan P, Latini F. The retroﬁtting
of an extended aeration process to optimise biological
nitrogen removal in liquid industrial wastes. Environ
Technol 1999;20:563–73.
4221
[25] Harrem.oes P, Sinkjaer O. Kinetic interpretation of
nitrogen removal in pilot scale experiments. Water Res
1995;29(3):899–905.
[26] Carrera J, Torrijos M, Baeza JA, Lafuente J, Vicent T.
Inhibition of nitriﬁcation by ﬂuoride in high-strength
ammonium wastewater in activated sludge. Process
Biochem 2003, in press.
[27] Hasselblad S, Hallin S. Intermittent addition of external
carbon to enhance denitriﬁcation in activated sludge.
Water Sci Technol 1998;37(9):227–33.
[28] Nyberg U, Andersson B, Aspergren H. Long-term
experiences with external carbon sources for nitrogen
removal. Water Sci Technol 1996;33(12):109–16.
[29] Bailey W, Tesfaye A, Dakita J, McGrath M, Daigger G,
Benjamin A, Sadick T. Large-scale nitrogen removal
demonstration at the Blue plains wastewater treatment
plant using post-denitriﬁcation with methanol. Water Sci
Technol 1998;38(1):79–86.
[30] Timmermans P, van Haute A. Denitriﬁcation with
methanol. Fundamental work of the growth and denitriﬁcation capacity of Hyphomicrobium sp. Water Res
1983;17(10):1249–55.
[31] Carrera J, Vicent T, Lafuente FJ. Inﬂuence of temperature
on denitriﬁcation of an industrial high-strength nitrogen
wastewater in a two-sludge system. Water SA
2003;29(1):11–6.