Introduction
Bios-sand filters (BSFs) are intermittent slow sand filters designed for household
use and hence called point-of-use (POU) water filtration systems, with
principal filtration mechanisms being physical, chemical and biological
(Murphy et al., 2010). The biological mechanisms take place at the top layer,
where a biological mat develops in the 50 to 100 mm of the media (CAWST,
2009). The biological layer acts both as a fine filter to remove small
colloidal particles, dissolved impurities and at the same time immobilizes
pathogens.
Although BSFs are now widely applied in the treatment of water
at household level, few studies have been conducted on the removal of
chemical contaminants. Current research in BSFs has mainly focused on the
removal of pathogenic organisms like Escherichia coli (E. coli) and suspended solids (Elliott et al.,
2008; Van Halem et al., 2009; Mwabi et al., 2012). One chemical of major
concern is nitrate-nitrogen contamination (NO3-N) on surface and ground
water as it poses serious health problems (Almasiri and Kaluarachchi, 2007).
Methaemoglobinemia in infancy is related to nitrate ingestion resulting in
low oxygen intake and consequently causing death (Craun et al., 1981;
Aslan and Cakici, 2007). Furthermore, presence of nitrates in drinking water
results in the formation of nitrosomines in the stomach, which are
carcinogenic (Shuval and Gruener, 1977; Aslan and Cakici, 2007; Speijers and Fawell,
2011). Nitrate poisoning has been reported in livestock when concentrations
exceeded 100 mgL-1 (Tredoux et al., 2000) and other problems related to
nitrate in drinking water are well documented in the literature (Moraes, 1996;
Fan and Steinberg, 1996; Lin et al., 2002; Forman, 2004).
Main sources of NO3-N on surface waters and groundwater aquifers
include use of agricultural fertilizers, animal waste disposal and wastewater
effluents from conventional and on-site sanitation facilities. Water supply
from high-nitrate concentration environments needs some form of treatment or
dilution with low-nitrate content water. The current design of conventional
BSFs has been proved to be poor in the removal of nitrates
(Heather et al., 2010; Mahlangu et al., 2011; Kennedy et al., 2012).
Physical and chemical methods such as ion-exchange, reverse osmosis,
electro-dialysis, distillation, nanofiltration and activated carbon have
been applied in the removal of nitrates from drinking water supplies
(Schoeman and Steyn, 2003; Shaharudin et al., 2017). These methods are
relatively expensive and show poor selectivity for nitrate removal with
generation of brine, which is difficult to dispose of (Mohseni-Bandpi et al., 2013).
Hence there is need to explore alternative technologies like biological
denitrification which has been proved to be efficient in complete nitrate
elimination and has the advantage of producing a harmless by-product
(N2). The pathway for nitrate removal by heterotrophic bacteria
is nitrate → nitrite → nitric oxide → nitrous oxide → gaseous diatomic nitrogen:
NO3-→NO2-→NO→N2O→N2↑.
The biological denitrification technology is based on the conventional
theory that carbon is the limiting factor in the efficiency of biological
denitrification. Heterotrophs utilize carbon from organic compounds like
sugars, organic acids and amino acids as source of electrons rather than
from inorganic compounds like carbon dioxide as is the case in autotrophic
denitrification (Mohseni-Bandpi et al., 2013). Although autotrophic nitrate
removal has the advantage of not requiring an organic carbon source, it is
associated with slow growth rate of autotrophic bacteria and low nitrate
removal rate (Mohseni-Bandpi et al., 2013).
Few studies have been conducted on the effectiveness of the BSFs in
the removal of nitrates. In a study conducted in rural Cambodia by Heather
et al. (2010), it was revealed that there was simultaneous nitrification and
denitrification occurring in the BSFs. However, about 85 % of
the biofilters under the study did not meet the WHO guideline for
NO3-N
in the treated effluent. The study showed that denitrification was
predominant when the inflow into the filter was from surface water, which
could be due to the high organic carbon content. Kennedy et al. (2012)
studied the effects of hydraulic loading on removal of nitrates in BSFs and the overall nitrate removal efficiency was low (16 %).
Mahlangu et al. (2011) established that the conventional BSFs and the
modified BSFs of zeolites (clinoptilote) have relatively low removal rates of
nitrates (37 %). In the same study, other types of biofilters which
include ceramic candle and bucket filters had poor rates of removal of nitrates
ranging from 18 to 37 % (Mahlangu et al., 2011). On certain
occasions, the effluent concentration of NO3-N was even higher than in the
unfiltered water, possibly due to desorption of previously adsorbed
nitrates and nitrification. Research has also revealed that heterotrophic
nitrifying microorganisms are key players in the nitrogen cycle and
can increase the effluent concentration of NO3-N through
cell lysis (Masahito et al., 2008).
Most sources of drinking water lack sufficient quantities of organic carbon
for cell growth as well as for energy source for the heterotrophic bacteria
(Mohseni-Bandpi et al., 2013). The organic carbon acts as both a source of
cellular material for biological respiration and electron donor for
dissimilatory nitrate reduction. Waters with low carbon content require an
external carbon source for denitrification to take place under anoxic
conditions and nitrate is converted to gaseous diatomic nitrogen.
A variety of external carbon sources like sucrose, ethanol, methanol and
acetic acid have been applied in conventional slow sand filters to aid
heterotrophic denitrification at C / N ratios ranging from 1 to 2.5 (Callado,
2001; Gomez et al., 2000; Aslan and Cakici, 2007). The studies have shown
considerable improvement levels in the denitrification process due to the
recorded high nitrate removal efficiencies of about 90 % (Green et al.,
1994). Gomez et al. (2000) assayed the influence of sucrose, ethanol,
methanol and ethyl alcohol in nitrate reductase in contaminated groundwater
and showed very high removal rates with effluent concentrations ranging from 0 to
5 mgL-1. Aslan and Cakici (2007) reported removal rate of 94 % for nitrate
in slow sand filters when acetic acid was used as a carbon source. Methanol
is toxic due to some of the residual concentrations of carbonaceous
compounds found in the effluent and produces an excessive growth of biomass
(Stouthamer, 1992; Cherchi et al., 2009; Jensen and Darby, 2012). Sucrose and
glucose have a tendency to form a biomass which increases turbidity in the
final effluent. Acetic acid and ethanol are considered to be the most
suitable carbon sources for removal of nitrate and no limits have been set in
potable water (Ghararah, 1996). They are also cheaper, a concept inherent
in the use of bio-sand filtration technology.
However, heterotrophic denitrification has not been investigated in
BSFs except in the conventional slow sand filters. The aim of
this study was to investigate the removal of NO3-N in BSFs
with ethanol as a carbon source and to establish the optimum carbon-to-nitrate (C / N) ratio for microbial activity which achieves maximum removal
with minimum excess carbon in the effluent.
Materials and methods
Two BSFs were investigated at household level: one with an
external carbon source (BSFC) to enhance the denitrification process at C / N
ratios of 1.1 and 1.8, and the other one without a carbon source (BSFW).
Figure 1 shows the schematic diagram of the BSFs which were
constructed. These two ratios were selected based on the optimum range of
carbon-to-nitrogen ratio which was established by Aslan and Cakici (2007), Gomez
et al. (2000) and Callado (2001) for denitrification in slow sand
filtration, which ranged from 1.08 to 2.5. The two BSFs were
dosed with known concentrations of ammonium nitrate, which was the source of
nitrate.
Filter construction
The two BSFs were constructed according to the Centre for
Affordable Water and Sanitation Technology guidelines (CAWST, 2009). Plastic
buckets 25 mL in volume were used and were packed with multi-media filter
material. The multi-media filter bed consisted of fine sand of 0.3 mm
diameter and 250 mm deep; sand of 0.95 mm diameter and 750 mm deep; gravel of
7 mm diameter and 50 mm deep. The South African National Standard (SANS 3001)
was used to determine the particle size and grading in order to achieve the
required particle size distribution of the filter media. Dewatering of the
filter between charges is avoided by a vertical discharge tube that rises
from 2 to 7 cm above the height of the filter media. The elevated outlet allows
the media to remain saturated after a charge has been filtered and when
water is no longer flowing from the outlet (Fig. 1). The design parameters
of the filter are summarized in Table 1.
Schematic representation of the BSF (dimensions in mm).
The direction of filtration flow is from top to bottom of filter bed.
Summary of the design values used for the two filters (BSFW and
BSFC).
Design parameter
Unit
Recommended
Reference
Applied
value
value
Media depth
m
0.3–0.5
CAWST (2009); Kubare and Haarhoff (2010)
0.3
Supernatant depth
mm
50
Lukacs (2002); Duke et al. (2006); CAWST (2009)
50
Surface area
m2
0.06
CAWST (2009)
0.071
Effective size
mm
0.15–0.40
CAWST (2009), Manz et al. (1993)
0.35
Coefficient of uniformity
–
1.5 to 3
Elliot et al. (2008); Manz et al. (1993);
2.64
Filtration velocity (in clean filter bed)
mh-1
0.10 to 0.6
Kubare and Haarhoff (2010); Elliot et al. (2008);
0.17–0.63
Inflow rate
m3h-1
0.03 to 0.04
CAWST (2009)
0.04
The filtration cycle of a biofilter is made up of resting time (6–24 h) and
a maximum filtration time of about 2 h (Fewster et al., 2004). The
biological treatment occurs during the resting time and after this period
the filter bed is drained. In this study the raw surface water or untreated
river water was fed into the filter once a day and the resting time and
filtration time were 24 and 2 h respectively. The filtered water was
collected in a 5 L vessel for laboratory analysis. The average inflow
rate was measured from noting the start time of filtration and the time
periods at which the level of the water in the receiving vessel changed by 1 L.
The superficial velocity (vs) is related to the surface area of the
filter and is normally used in filtration computations and is also
equivalent to the hydraulic surface loading divided by the surface area of
the filter. For BSFs, the inflow rate is not constant since the water is only
poured once for a filter cycle and hence the infiltration velocity decreases
with time from the start to end of cycle.
Nitrate and carbon source dosage
An influent nitrate concentration of 25 mgL-1 was selected based on the
guideline value of 11mgL-1 in potable water and the average reported values
of nitrates in surface- and groundwaters, which range between 0 and 18 mgL-1
(WHO, 2011). The raw river water had a low background concentration of
nitrate ranging from 0.39 to 1.15 mgL-1. However, large parts of southern
Africa have nitrate values often exceeding 50 mgL-1 (Tredoux, 2004). A stock
solution of ammonium nitrate (NH4NO3) of concentration of
190 gL-1
was dosed to both filters (BSFW and BSFC), and to achieve a dose of 25 mgL-1 in
the 25 L filter volume, 3.33 mL of the stock solution was required. The
ethanol was applied only to BSFC at C / N ratios of 1.1 and 1.8. With a molar
mass of 46 gmol-1 of ethanol (C2H5OH) the carbon equivalent in the
ethanol was 24 gmol-1 (52.2 %). Therefore, at a nitrate dose of 25 mgL-1 and
C / N ratio of 1.1 the dosage of carbon as ethanol in a 25 L BSF was
7.45 mL of carbon as ethanol. Similarly, at C / N ratio of 1.8, the required
dose of carbon as ethanol was 12.1 mL.
The surface loading of NO3-N was calculated by multiply the
concentration of nitrate with the superficial velocity (gm-2d) and the
denitrification rate was computed as
Rdn=1t(Cin-Cout),
where
Rdn is denitrification rate (ML-3T), Cin is influent nitrate (ML-3) and
Cout is effluent nitrate (ML-3).
Filter maturation
The denitrification process in BSFs is biological and takes place
under a fixed film growth process whereby the bacteria develop on the
surface of the sand media. For the smooth operation of the BSF,
the water level was maintained at 50 mm above the fine sand. The maturation
period for the full development of the biological layer and acclimatizing of
the microorganisms to ethanol and NO3-N environment was 3 weeks. The
biological layer typically takes 20 to 30 days to develop to maturity in a
new filter depending on the quality of the inlet water (CAWST, 2009;
Mahlangu et al., 2011). The operating temperatures of the filters varied between
19 and 20 ∘C and were not controlled, in order to simulate the
actual operating conditions of a BSF at household level.
Sample collection and analysis
Sampling bottles were washed with distilled water before and after sampling.
The samples were collected at the inlet and outlet of the two BSFs in 500 mL Erlenmeyer flasks and stored in a refrigerator at 4 ∘C
and analysed within 1 h. The frequency of sample collection was once a
week after the 12 h resting time.
The pH and dissolved oxygen (DO) were measured using a pH meter, model HACH
HQ30D (FLEXI model). The instrument was calibrated and measurements
conducted in accordance with the standard method. The nitrate was measured
by Spectroquant nitrate photometrical test method using Merck
spectrophotometer PHARO100 and the results were reported as NO3-N in
mgL-1. The carbon source which was ethanol was measured as chemical oxygen
demand (COD) by the MERCK Spectroquant TR 320 digester (Spectroquant COD
cell test method). The samples were digested in tubes containing a mixture
of chromic and sulfuric acid with silver sulfate as a catalyst. After
digestion samples were cooled and read on the Spectroquant PHARO100
spectrophotometer. The COD test was carried out mainly to determine the
amount of ethanol as a carbon source in the source water before and after
the filtration process.
Results and discussions
Flow rates
Initial flow rates in the control filter BSFW started from 0.04 m3h-1 and
declined to 0.03 m3h-1 by the end of the experiment. In BSFC which
received the carbon source the flow rate reduced from 0.04 to
0.01 m3h-1 (Fig. 2). The reduction in flow rates was comparable to
studies conducted on BSFs by Kubare and Haarhoff (2010) and Kennedy et al. (2012).
The decline in the filtration rate was due to filter clogging and was
substantial when the biological layer was fully mature. The reduction in the
flow rate was more pronounced in the filter dosed with an external carbon
source (BSFC) compared to one without carbon (BSFW). Therefore, there was
more growth of the biomass in the biofilter with an external carbon source
due to the favourable environment conducive for growth of heterotrophic
bacteria. Conventional surface cleaning will not remove the biomass at the
bottom layers. Consequently, a household would require more filters to meet
the daily water demand as well as increasing the resting period in BSFC to
reduce excessive growth of biomass. Overall, the filtration velocity ranged
from 0.17 to 0.63 mh-1 and typical filtration rates for BSF range from
0.16 to 1.1 mh-1 (Elliot et al., 2008; Kubare and Haarhoff, 2010).
Changes in pH and DO
The pH and DO are important physicochemical parameters in the removal of
nitrates in BSFs. There was no significant change in the pH of the influent
and effluent water for both filters (BSFW and BSFC). Overall, there was a
slight decrease in pH from 8.6 to 6.8 and such a pH range would favour the
denitrification process since maximum denitrification rates are achieved
at pH range of 7 to 8.5 (Wang et al., 1995), whereas pH values smaller
than 6 and larger than 8.5 would result in a sharp decrease in the
denitrification activities (Drtil et al., 1995). The slight decrease in pH
could be due to nitrification and aerobic respiration at the top layer of
the filter due to availability of oxygen and this phenomenon was also
confirmed by Heather et al. (2010) and Mangoua-Allali et al. (2012). Nitrification is
obligatorily coupled to oxygen consumption and has an effect on the decrease
in alkalinity. Such a decrease in alkalinity might cause a decrease in pH
because an acidic nitrite formation results in a drop in pH. Thus, if the
buffer capacity of the system is weak, the pH might drop well below 6.7
(Habboub, 2007).
Variation in flow rates in the filters with and without carbon
source.
However, pH may increase during denitrification because the reduction of
nitrate to gaseous nitrogen with organic substrate as an electron donor
results in the production of carbon dioxide and oxygen hydroxide (OH-),
which may react to form a bicarbonate (HCO3-) and carbonate
(CO32-) (Drtil et al., 1995; Wang et al., 1995). With regard to
water quality guidelines, the pH values were within the acceptable South
African guideline limits of 5.0 to 9.7 (SANS 241-2, 2015).
The overall reduction of DO in the filter with an external carbon source was
65 % with average inflow and outflow concentrations of 8.23 and
2.94 mgL-1 respectively.
DO concentration is influenced by a number of factors
including water temperature, organic matter, salinity and atmospheric
pressure. The operating temperature of the filters was between 19 and
20 ∘C and the measured DO values are typical at such temperatures.
Furthermore, the water which was used was raw river water, and the DO can
range between 0 and 18 mgL-1 in such waters depending on the level of pollution.
However, the reduction in dissolved oxygen was less in the filter without an
external carbon source (50 %). The reduction in the DO is due to the
oxygen demand by aerobic and nitrifying bacteria at the top layer of the
filter bed.
Nitrate removal rates
The nitrate removal mechanisms during heterotrophic denitrification are
bacterial respiration and bacterial synthesis (Mohseni-Bandpi et al., 2013).
The denitrification will take place at the bottom of the filter bed where
there is less oxygen (anoxic conditions). William and Beresford
(1998)
concluded that nitrification and denitrification happen simultaneously in
zones where there are short distances between the aerobic and anaerobic
zones. The same scenario is depicted in BSFs due to the short
filtration length of approximately 0.3–0.5 m (Elliot et al., 2008; CAWST, 2009).
Heterotrophic bacteria need organic carbon as the electron donor and as the
source of carbon, whilst getting their oxygen by removing bound oxygen from
nitrate (NO3-) which is in the water being treated. The nitrate
acts as the electron acceptor. As a result of this process, the removal rate
of nitrates in the filter without external carbon source (BSFW) was
30 % ± 0.04 (Table 2) and Mahlangu et al. (2011) reported a rate of 37 %
in similar filters. In the filter with an external carbon source (BSFC) the
nitrate removal rate was 44 % ± 0.03 at C / N ratio of 1.1 and
53 % ± 0.03 at C / N ratio of 1.8. Overall, the nitrate removal rate was
higher with the use of an external carbon source at higher C / N ratio of 1.8.
The reason for this is that carbon is the limiting factor in denitrification
since heterotrophic bacteria need organic carbon as the electron donor and
as the source of carbon. Therefore, a higher carbon content will result in a
higher nitrate removal rate. However, the effluent nitrate concentration of
between 16 and 19 mgL-1 was still above the recommended guideline values for
potable water.
Nitrate removal efficiency at C / N = 1.1; C / N = 1.8 and at influent
nitrate concentration of 25 mgL-1.
BSFW (without external carbon)
BSFC at C / N=1.1
BSFC at C / N=1.8
Sampling interval
Effluent nitrate
Removal
Effluent nitrate
Removal
Effluent nitrate
Removal
(days)
(mgL-1)
efficiency (%)
(mgL-1)
efficiency (%)
(mgL-1)
efficiency (%)
1
19.21
23
15.55
38
14.21
43
2
19.00
24
14.95
40
13.08
48
5
18.75
25
14.90
40
13.01
48
7
16.25
35
14.81
41
12.85
49
9
16.50
34
14.74
41
12.81
49
12
17.00
32
14.61
42
12.75
49
14
17.50
30
14.55
42
12.73
49
17
16.00
36
14.50
44
12.70
49
20
16.32
35
14.50
42
11.75
53
22
16.42
34
14.50
42
11.60
54
24
16.30
35
13.50
46
11.50
54
27
15.64
37
13.00
48
11.75
53
29
15.73
37
12.65
49
11.90
52
The failure to achieve effluent nitrate guideline values even though pH was
optimum could be due to high DO. Optimum denitrification occurs under
anoxic conditions when oxygen levels are depleted (low redox) and nitrate
becomes the primary oxygen source for heterotrophic bacteria. In general, it
has been observed that a DO concentration of more than 0.2 mgL-1
reduces the rate of denitrification significantly (Jorgensen and Sorensen,
1988). High levels of DO were recorded ranging between 2.9
and 8.2 mgL-1 – higher than the optimum values for
denitrification.
Reducing the DO concentration in a BSF will enhance the nitrate
removal efficiency but will compromise the aerobic microbial activity at the
top layer. A feasible alternative would be to increase the filter depth so as
to create an anoxic zone at the bottom or to increase the resting period of
the filter. BSFs are designed with a filtration time of
2 h
and resting period of 12 to 24 h (CAWST, 2009; Elliott et al., 2008).
The resting time provides the contact time for microbial removal and
denitrification processes and thus a long resting time is desirable
from that perspective. However, too long a resting period may reduce the
viability of the biological layer because the survival of the microorganisms
relies on the periodic inflow of source water for nutrients (Baumgartner et
al., 2007). Additionally, too long a resting period will reduce the water
production rate and thus fail to satisfy household water requirements.
Therefore, careful selection of the resting period is vital in order to
balance these competing objectives. In this study a resting time of
12 h
was used and nitrate concentrations measured during this period showed a
rapid removal rate during the first 1.5 h and no significant removal
thereafter (Fig. 3). Therefore, increasing the resting period by more than
12 h will not have any significant effect on nitrate removal. Results for the entire operational period indicate low removal at the
beginning (40 %). Thereafter the rate increased to 53 %. This
illustrates the importance of maturation period. The variation in nitrate
concentrations for the entire operational period is shown in Fig. 4.
The average denitrification rates for BSFW and BSFC were
3.66 and 5.44 g NO3-N m-3 days respectively and
these rates are lower than those reported by Aslan and Cakici (2007) in slow
sand filters (ranging between 8.1 and 29.2 g NO3–N m-3days at
filtration rates between 0.015 and 0.06 mh-1).
Reduction of nitrate relative to resting period in the filter with
an external carbon source. Values of the nitrate are the average of the C / N
ratio of 1.1 and 1.8.
Variation in nitrate concentrations for the entire operational
period.
Residual COD in effluent
The residual ethanol measured as COD in filters with an external carbon
source varied between 25 mg and 36 mgL-1. Overall, the removal efficiency of COD
at C / N ratio of 1.1 and 1.8 was 89 and 91 % respectively (Table 3).
COD removal efficiency at C / N =1 and C / N=1.8 and influent COD
of 233.52 and 382.12 mgL-1 respectively.
Sampling interval
Effluent COD
COD removal
Effluent COD
COD removal
(days)
(mgL-1)
efficiency (%)
(mgL-1)
efficiency (%)
C / N=1.1
C / N=1.8
1
26.85
88.50
34.62
90.94
2
25.30
89.17
35.17
90.80
5
25.61
89.03
35.98
90.58
7
23.98
89.73
34.48
90.98
9
24.77
89.39
35.96
90.59
12
25.10
89.25
34.84
90.88
14
26.36
88.71
34.46
90.98
17
24.67
89.44
35.54
90.70
20
26.46
88.67
36.10
90.55
22
26.70
88.57
35.86
90.62
24
26.55
88.63
35.42
90.73
27
26.48
88.66
35.40
90.74
29
26.22
88.77
35.18
90.79
There was rapid COD removal in the first 2 h. The rate stabilizes as the
resting period increases and hence there is no significant benefit in
longer resting periods. The same trend is seen for nitrate removal,
which suggests that the denitrification process takes place in the first 2 h
when the COD is utilized in the process. However, the COD
concentrations in the effluent were higher than the guideline values, and
such high levels of COD concentrations can be toxic to human health and
increase disinfection by-product formation potential. This represents a major
health challenge in the use of an external carbon source for the removal of
nitrates in potable water and there is a need to explore post-treatment
methods to remove the residual carbon in BSFs.
Reduction of COD relative to resting period in the filter with an
external carbon source.