Remediation of nitrate-contaminated water by solid-phase denitrification process-a review.

Environ Sci Pollut Res DOI 10.1007/s11356-015-4334-9

REVIEW ARTICLE

Remediation of nitrate-contaminated water by solid-phase denitrification process—a review Vaishali Ashok & Subrata Hait

Received: 3 December 2014 / Accepted: 6 March 2015 # Springer-Verlag Berlin Heidelberg 2015

Abstract The paper presents a compilation of various autotrophic and heterotrophic ways of solid-phase denitrification. It covers a complete understanding of various pathways followed during denitrification process. The paper gives a brief review on various governing factors on which the process depends. It focuses mainly on the solid-phase denitrification process, its applicability, efficiency, and disadvantages associated. It presents a critical review on various methodologies associated with denitrification process reported in past years. A comparative study has also been carried out to have a better understanding of advantages and disadvantages of a particular method. We summarize the various organic and inorganic substances and various techniques that have been used for enhancing denitrification process and suggest possible gaps in the research areas whi'ch are worthy of future research. Keywords Denitrification . Solid phase . Autotrophic . Heterotrophic

Introduction Nitrate is the world’s most common pollutant in the terrestrial ecosystem (Trudell et al. 1986; Schwientek et al. 2008). Nitrate contamination has become a serious environmental Responsible editor: Bingcai Pan V. Ashok (*) Department of Civil Engineering, Indian Institute of Technology (IIT) Kanpur, Kanpur, India e-mail: [email protected] S. Hait Department of Civil and Environmental Engineering, Indian Institute of Technology (IIT) Patna, Patna, India

concern for both surface water and groundwater ecosystems (Rivett et al. 2008; Moraes and Souza 2012; Gong et al. 2013; Jiang et al. 2013; Zhu et al. 2013). In past few years, the nitrogen content in groundwater has tremendously increased due to various anthropogenic activities like application of animal manure and intensive use of fertilizers to agricultural land (Puckett et al. 2002; Ibendahl and Fleming 2007; Schwientek et al. 2008; Torrento et al. 2010). Other factors that contribute toward nitrate contamination of groundwater are atmospheric deposition (Zhu et al. 2013), percolation from leaking sewers, discharge from septic tanks, and application of sewage sludge and animal wastes on agricultural land (Wakida and Lerner 2005). Increase in nitrate concentrations can cause adverse effects on different ecosystems with varying degree of exposure levels (Table 1). It may lead to eutrophication, toxic algal blooms, habitat deterioration in river, lakes, and various coastal regions, hypoxia, and increased N2O emissions in atmosphere (Galloway et al. 2003; Warneke et al. 2011a, b; Zhou et al. 2011). Use of such poor quality of water in irrigation can cause reduced crop production (Azizullah et al. 2011). Consumption of nitrate-contaminated drinking water may cause risk to human health such as shortness of breath and cancer (Weyer et al. 2001; Schwientek et al. 2008). When nitrate is consumed by infants in huge quantities, then on their skin appears a bluish tint due to lack of oxygen called methamoglobinemia or blue baby syndrome (Ward et al. 2005; Torrento et al. 2011). High nitrate concentrations in drinking water can also cause hypertension, birth defects, and abortions. In animals, it may lead to muscular weakness and abdominal pains (Naik and Setty 2011). European Union Council Directive (98/83/EC) has set a standard of 50 mg NO3−/l in drinking water. USEPA has set limit of 10 mg/l of nitrate nitrogen, and Beijing, since July 2012, has enforced total nitrogen levels to be 10 mg/l (BMEPB and BMAQTS 2012) (Li et al. 2013).

Environ Sci Pollut Res Table 1 Properties of different ecosystems on nitrogen accumulation and transfer potential (adapted from Galloway et al. 2003) Ecosystem

Accumulation potential

Transfer potential

Atmosphere Low Very high Wetlands, Low Very high streams, lakes, rivers Agro-ecosystems Low to moderate Very high Marine coastal regions Groundwater Forests Grasslands

Low to moderate Moderate Moderate High High

N2 production potential None Moderate to high Low to moderate High

Moderate Moderate Moderate to high Low Moderate to high Low

Despite of the stringent laws, contamination of groundwater with nitrate has been reported worldwide with varying degree of contamination levels. The groundwater in Osona area (Spain) is affected by nitrate contamination due to excessive application of pig fertilizer in fields (Torrento et al. 2011). Nolan et al. in 1997 reported 25 % of the wells in USA under high risk condition. Cordy et al. 1995 found median concentration of nitrate as 1.4 mM in Central Arizona basins. Zhang et al. (1996) reported a 140,000-km2 area in northern China effected with nitrate contamination with limit exceeding 50 mg/l, which exceeds permissible limit of drinking water. Some areas among them show extremely high concentrations of 300 mg/l of nitrate in ground and drinking water near farmer’s yard, center of small cities, etc. Suthar et al. (2009) have investigated 21 villages of Rajasthan, India, for nitrate contamination of groundwater and found average nitrate levels of 60.6±33.6 mg/l, showing high pollution levels due to anthropogenic activities. The groundwater pollution is more severe in China where nitrate nitrogen in some rural areas exceeds 130 mg/l (Wan et al. 2009). Devic et al. (2014) reported high groundwater pollution in Serbia and water quality in category III/IV representing contaminated water not suitable for human consumption. Comparative studies of nitrate contamination in groundwater of major cities of the world have been presented by Wakida and Lerner (2005). Nitrate remediation from groundwater is a difficult task as the sources are diffuse, control measures on different land use patterns with different rates of loading are problematic, and natural attenuation process is very slow. The study focuses on various autotrophic and heterotrophic removal technologies and substrates which are being progressively being used for the nitrate removal from water resources. It also addresses their drawback and challenges associated with it while applying to a field scale. A brief description about the denitrification process and factors on which it depends is described firstly and then the removal methodologies in the coming sections.

Solid-phase denitrification process The most significant natural process of nitrate reduction is denitrification, where certain denitrifying microorganisms use carbon sources as electron donor and nitrate as electron acceptor and reduces it to dinitrogen gas under a given set of conditions (Knowles 1982; Zumft 1997; Alvarez et al. 2007; Torrento et al. 2010). Denitrification in a natural system is a very slow process. The term Bsolid-phase denitrification^ came into existence in recent years, where solid substrates were used as a constant source of denitrification and also provide platform for microbial biomass development. Solidphase denitrification process can be achieved by both heterotrophic and autotrophic ways, depending upon whether the energy derived by microorganisms is from organic sources or from inorganic sources. The common pathway of heterotrophic denitrification in wetlands is shown in Fig. 1 (Xu et al. 2013). Nitric oxide (NO) and nitrous oxide (N2O) are the intermediate species during denitrification process but, under a given set of favorable conditions, get rapidly transformed to nitrogen gas (N2). N2O is often used as an indicator of denitrification occurrence (Yoshinari et al. 1977; Mathiesen 1998; Jahangir et al. 2013). N2O concentrations found in shallow ground waters were threefold higher than that in water at equilibrium with atmosphere. Nitrous oxide concentrations were found to be highest near or at the water table and it reduces with depth. This indicates that nitrous oxide loading might have been derived from nitrification-denitrification process in vadose zones and was being transferred to groundwater table via recharge (Spalding and Parrott 1994). NO and N2O formation is not desired as they contribute to greenhouse effect, acid rain, and ground-level ozone formation. Denitrifying microorganisms are ubiquitous in nature, whether be it surface water (Bernot et al. 2003), groundwater (Jacinthe et al. 1998; Vogel et al. 1981; DeSimone and Howes 1998; Smith and Duff 1988; Peterson et al. 2013; Mohamed et al. 2003), soil, or at great depths below ground levels (Morris et al. 1988). Denitrifying bacteria are mostly facultative anaerobic heterotrophs and obtain their energy from the oxidation of organic substances (Torrento et al. 2011). Autotrophic denitrifiers generally derive their energy from sunlight; however, in the absence of oxygen, some autotrophic denitrifiers derive their energy from inorganic sources also. (Rivett et al. 2008). Many isotope-based techniques have been employed in different parts of the world to identify denitrification rates in groundwater and streams (Fukada et al. 2003; Veraart et al. 2013). During the process, microorganisms were reported to be inclined toward lighter isotopes of 14 N and 16O than 15 N and 18O in the approximate ratio of 2:1 (Bottcher et al. 1990; Aravena and Robertson 1998; Cey et al. 1999; Mengis et al., Lehmann et al., 2003; Schwientek et al. 2008). Some physico-chemical technologies which have been functioning parallel with biological denitrification processes

Environ Sci Pollut Res Fig. 1 Nitrogen pathway for heterotrophic denitrification in wetland areas

Denitrificaon

Ammonificaon

Organic N

NH4+

NO2-

NO3-

NO2-

NO

N2O

N2

Nitrificaon

are ion exchange (removal through adsorption of ammonium or nitrate ions through charged resin columns) (Chen et al. 2002; Samatya et al. 2006), electrodialysis (removal by passing water through ion-selective semi-permeable membranes with applied potential difference between the two membranes) (Cheikh et al. 2013), reverse osmosis (removal by forcing the water to pass through a semi-permeable membrane and leave nitrate and other ions behind), etc. (Kapoor and Viraraghavan 1997). Disadvantages associated with these technologies are brine generation, toxic by-products, high cost, and poor selectivity (Alvarez et al. 2007; Wan et al. 2009). Conditions favorable for solid-phase denitrification process are discussed in the following sections such as (a) availability of nitrate and electron donor species, (b) low dissolved oxygen concentration (1–2 mg/l) (Cey et al. 1999; Robertson and Merkley 2009), and (c) other favorable conditions (like pH, temperature, presence of other species, trace elements, or heavy metals (Xu et al. 2009; Torrento et al. 2010)).

Nitrate concentration Input nitrate concentration plays a major role in determining the denitrification efficiency of a system. The nitrate removal efficiency increases with decrease in initial nitrate concentration and grain size (Torrento et al. 2011). Greater influent nutrient concentration will require greater hydraulic retention time (HRT) for complete removal (Zhou et al. 2011). Also, high initial concentration can affect denitrification process by inhibiting the production of N2 gas from N2O reduction (Blackmer and Bremner 1978).

Oxygen concentration Oxygen is the preferable electron acceptor for the oxidation of organic carbon sources (Fig. 2). This highlights the fact that denitrification is predominantly an anaerobic process. Denitrification generally occurs at dissolved oxygen (DO) content less than 2 mg/l. At lower depth, it may reach a dissolved oxygen concentration of less than 1 mg/l (Fukada et al. 2003)

Adaptability of various microorganism Denitrification is also affected by different bacterial populations. Denitrifying bacteria are ubiquitous in subsurface environment. These microorganisms are generally heterotrophic in nature and use carbon source as the electron donor species. Minority of bacteria also show chemolithotropic denitrification and use inorganic compounds, for instance, ferrous iron, sulfur compounds, inorganic carbon, hydrogen, iron, or alloyed nanoparticles and uranium as electron donors (Straub et al. 1996; Zumft 1997; Beller 2005). Lag time refers to the adaptability of a particular or group of species to adapt to new set of conditions. A lag phase of some days or hours have been generally encountered when denitrification process accompanied is with microbial population (Fellows et al. 2011). The lag time of different species varies with type of water, presence of trace metals, cell growth rate of that particular species, temperature, etc. Vogel in 1996 suggested that in the absence of appropriate denitrifying bacteria, specialized bacterial species should be introduced for fast removal of nitrate (Torrento et al. 2011). Selection of appropriate denitrifying bacterial strain with steady biofilm formation ability enhances the proper functioning of reactor and simultaneously leads to a high-quality treated water (Moreno et al. 2005). pH pH plays a vital role in sludge granulation and sludge flocculation. Less than 8 pH supports faster sludge granulation (Bhuvanesh et al. 2013). Denitrifying bacteria has a convenient pH range of 6–8.5. pH values higher or lower than this range can cause sharp deterioration in denitrification potential (Bhuvanesh et al. 2013). Since pH influences the enzyme activity of bacteria, it plays a major role in nitrite accumulation in the denitrification reactors (Zhou et al. 2007). Temperature Various studies have been conducted to establish the relationship between temperature and denitrification rates in diverse ecosystems (Bouletreau et al. 2012). High temperature favors better denitrification efficiency than low temperature. Zhou

Environ Sci Pollut Res Fig. 2 Various electron acceptors in saturated zones for the oxidation of organic carbon source in a thermodynamic sequence (Adapted from Rivett et al. 2008)

et al. (2011) reported that low temperature is unfavorable for the start-up of a denitrification reactor regardless of nutrient concentration in the reactor. With increased HRT, running autotrophic denitrification reactors can sustain low temperatures and achieve good removal efficiency. An experiment with starch/polyvinyl alcohol (PVA) interblend materials as solid carbon source for nitrate removal from secondary treated effluent was performed by Li et al. in 2013. They found that the denitrification rates increased from 2.2- to 3.6-fold higher when temperature was heightened by 23–30 °C. In a polycaprolactone filled packed bed reactor, the volumetric nitrogen removal rate declined by around 50 % when temperature was decreased by 5 °C. As with decrease in temperature, the efficiency of carbon source to hydrolyze decreases which results in reduced performance of denitrifying bacteria in biofilm, which may lead to reduced nitrogen removal efficiency (Canziani et al. 1999). HRT Hydraulic retention time (HRT) is a very significant factor to be considered during start-up of a reactor. In heterotrophic system, hydraulic retention time is a function of growth rate of microorganism, initial nitrate concentration, presence of other inhibitory species, and temperature. Zhou et al. in 2011 has recommended the following set of HRT for various nitrate contamination levels based on sulfur-limestone-based autotrophic denitrification system (Table 2). Grain size Large intra-granular space in the aquifer enhances the surface area and pore size for better microbial growth and metabolic activities (Blakey and Towler 1988). Large microbial populations are unable to grow in small fissures leading to a small available surface area for microbial growth which eventually leads to decrease in denitrification rates (Johnson et al. 1998). Surface area and particle size of elemental sulfur or iron

particles is considered to be an elemental factor on which chemolithotrophic denitrification rate depends (Germida and Janzen 1993; Watkinson and Blair 1993; Moon et al. 2006; Torrento et al. 2010). Electron donor availability Most critical limiting factor of groundwater denitrification system is the availability of organic carbon source (Zhang et al. 2012). The denitrification processes have been reported to be present in case of shallow groundwater with depth of 2– 3 m, because of the absence of organic carbon sources (Starr and Gillham 1993). On contrary, Alday et al. (2014) suggested that organic carbon can be transported to the deeper sections of the aquifer by density-driven flow. Yet, in absence of organic carbon as electron donor, sulfur and reduced iron were reported as electron donor species necessary for denitrification (Fig. 2) (Bottcher et al. 1990; Pauwels et al. 2000; Beller et al. 2004; Korom et al. 2005). Figure 2 illustrates the preferential order of common electron acceptor species based on their Gibbs free energy in a saturated region. Organic carbon has an inclination to get oxidized most preferably with the available electron acceptor species which can provide maximum energy to the microorganisms. In aerobic conditions, oxygen is utilized first by aerobic bacteria to oxidize organic carbon source. With oxygen depletion, simultaneously, other species based on their availability and free energy will act as electron acceptors. In the Table 2 Recommended HRT for various contaminated water systems (Zhou et al. 2011) Type of water

Nitrate/nitrite concentration

HRT (h)

Eutrophicated surface water

Nitrate 70 mg N/l

>6

Environ Sci Pollut Res

presence of common electron donors in the saturated ecosystem, oxidation of organic carbon is initiated with aerobic oxidation followed by denitrification and manganese, iron, and sulfate reduction.

of organic carbon sources has been compared based on their denitrifying potential in Table 3.

Presence of other trace materials

In the absence of other simpler carbon sources, heterotrophic bacteria can feed on dead and lysed cells as their energy and carbon source (Koenig et al. 2005; Torrento et al. 2011). Some studies report adequate denitrification potential in riparian zones as plenty of electron donor species are available such as dissolved organic carbon from plant root, living organism, and labile soil (Cey et al. 1999; Mengis et al. 1999; Vidon and Hill 2004; Parmentier et al. 2014). Soil also plays a salient role in wastewater treatment (Wakatsuki et al. 1993). Trois et al. (2010) reported complete denitrification of highly concentrated leachate nitrified sample with dead remains of plants and trees successfully. Trudell et al. (1986) reported spatial variability of denitrification potential occuring in a shallow groundwater system in vertical direction. The results were supported by Hill et al. (2000) in a riparian groundwater of Canadian stream in that with the variation of organic carbon source in groundwater, the denitrification rates will also vary simultaneously.

Presence of other trace metals can have advantages as well as disadvantages on the denitrification efficiency depending upon its interaction with other species and concentration. There are some micronutrients which are necessary for the microbial growth; however, some has inhibitory effects. Micronutrients such as B, Co, Cu, Fe, Mn, Mo, and Zn in addition to C, S, and P are needed by denitrifying bacteria to fulfill their metabolic requirements. These nutrients are generally available in most groundwaters in adequate concentrations to support microbial growth (Rivett et al. 2008). Also, presence of minerals such as sodium chloride and calcium ions plays a major role in sludge granulation which speeds up the nitrate removal process. There is almost a 10-fold escalation in the granule settling velocity as compared to the settling properties of activated sludge in the presence of sodium chloride and calcium ions (Bhuvanesh et al. 2013). Presence of some sulfur forms like sulfate, sulfide, and thiosulfate has reported to inhibit denitrification rates (Hiscock et al. 1991). Denitrification can also be inhibited by the existence of heavy metals, pesticides, and its derivatives (Bollag and Henneringer 1976; Bollag and Barabasz 1979). Bollag and Kurek 1980 reported pesticides such as N-formyl-4-chloro-o-toluidine, 4-chloro-o-toluidine and aniline intermediates to inhibit nitrate removal processes.

Solid-phase heterotrophic denitrification In solid-phase heterotrophic denitrification, denitrifying bacteria present in aquifers retrieve energy from solid organic sources. Heterotrophic bacteria are those which derive their energy from organic carbon substances such as dead and lysed cells, decayed matter, woodchips, and methanol. The processes which involve aeration process before anoxic zones mostly convert all available carbon sources into its oxides, and little is left available to denitrifying microorganisms in anoxic zones. Hence, there is a need of external carbon energy source for the nitrifying organisms. A variety of electron acceptor species in presence of organic carbon as electron donor, commonly present in the subsurface regions, has been presented in Fig. 2. Natural wetland soils are good for nitrate and phosphate removal as it is a natural source of plenty of organic carbon substances (Nichols 1983; Song et al. 2013). Mainly, two varieties of solid carbon sources have been investigated for denitrification from water sources. They are natural substances and synthetic polymers (Shen et al. 2013). A variety

Dead and lyzed cells, decayed organic matter, and compost

Woodchips Various organic materials are being checked for their denitrification potentials, of which wood is most commonly being used (Schipper and Vukovic 2001; Robertson and Merkley 2009) because of its low cost, availability, and high carbonto-nitrogen ratio (Gibert et al. 2008). Increasing the amount of wood, its surface area and temperature are positively correlated with the nitrate removal rates; however, wood size has no significant effect on the denitrification rates (Schmidt and Clark 2013). Warneke et al. (2011a, b) have compared four different approaches of measuring nitrate removal rates in denitrifying woodchip bioreactors serving the effluent from glasshouse. Softwoods showed a higher removal efficiency among hardwood, coniferous, mulch, willow, and compost (Gibert et al. 2008). A maximum removal efficiency of 99.7 % has been achieved with woodchips at a low hydraulic retention time of 2 days. Here, the reaction rates declines with the age of wood packing while the porosity of wood remains almost same with time (Leverenz et al. 2010). Sawdust and wheat straw Wheat straw and sawdust serves as a carbon source; in addition, it also provides a support for the growth of biomass. The immobilized biomass has been increasingly used for wastewater treatment as it reduces the size of bioreactor, and flow rates can be increased (Castilla et al. 2003; Cheikh et al. 2013). With time, the efficiency of the sawdust decreases and

130 ml

250 ml

Synthetic water

Synthetic

Drinking water

Wastewater from treatment plant

Synthetic

400 ml

Wheat straw

1 l, 5.7 l Softwood Hardwood Coniferous Mulch Willow Compost Leaves Native soil 30 l Maize cobs Green waste Wheat straw Softwood Hardwood 7600 l Woodchips

Groundwater

Lodgepole pine wood chips (LPW) Cardboard (CB) Lodgepole pine wood needles (LPN) Barley straw (BS) Soil

80

98.7 98.7 95.1 89.7 86.3 92.7 93.9 73.2 ~95 ~98 ~98 ~96 ~96 99.7

~74 ~28

~94 ~75

~70

100 100 100 100 100

70

Cellulose foam carriers enhance denitrification rates. Nitrate removing rates increases when cellulose is used in combination with mixed bacterial cultures Gracilaria verrucosa shows highest nitrate removal efficiency. Yielded highest nitrogen production. Protein yield was also high

Remarks

2 days

43.9 h 34 h 54.3 h 33–40 h 37.5 2 days

Cameron and Schipper (2010)

Gibert et al. (2008)

Healy et al. (2012)

Trois et al. (2010)

Ovez (2006)

Kambe et al. (2005)

Reference

Nitrate removal rates decline with temperature decrease. Leverenz et al. (2010) The reaction rates declines as woodchip packing ages. Porosity of woodchips remained almost same over study period. Woodchips showed an effective carbon source for nitrified septic tank effluent Addition of new wheat straw temporarily increases performance. Soares and Abeliovich Substrate also acts as bacterial mechanical support (1998)

Maize cobs show a 3–6.5 fold enhancement in nitrate removal rates over wooden media. Ammonia and BOD leaching found from maize cobs. Volume of bed with maize cobs was found to be 3 times lesser than based on wood chips

13 days 24 days 20 days 10–20 days No decline in performance of substrate during 3 months. 30–40 days No methane detection. Considerable COD production from both substrates 8–22 days PO4-P release in effluent was highest for LPW, followed by CB, LPN and BS. PO4-P decreased in all media, but outlasts above environmental thresholds. During steady state measurements COD release was highest for LPN. Green house gas emissions (particularly CH4 and N2O) were detected in all media. Greenhouse gas emissions were highest for soil control 135 days Nitrate removal greater than 95 % efficiency is found in almost all organic substrates. Nitrite and ammonia production. Of all substrates tested, softwood has highest denitrification rates. Compost adds extra leachable nitrogen along with ammonia formation

200 days

Influent Removal HRT nitrate-N efficiency conc. (mg/l) (%)

Comparison of denitrifying potential of various organic substrates

Type of water

Table 3


Sawdust

Sawdust, woodchip

–

–

250 ml

1.75 l

99

>96

80

–

100

>90

>99

89–95 85–94 >99

78 95 100

95

Remarks

With decrease in water table, denitrification rates decreased and increased with increase in water table. Treatment wall is able to remove nitrate for more than 2.5 years from groundwater with good removal efficiency – Dairy effluent shows better denitrification efficiency than other effluents. – Organic nitrogen, phosphorous and ammonium were not – removed 4 days BP shows higher nitrate removal efficiency and durability than wheat straw and saw dust. COD released by BP was low Ammonium accumulation 1.7 h No nitrite accumulation pH 7–8 is optimum for denitrification Higher denitrification rates at lower sections of the reactor ~14 h No nitrite accumulation Replacement of corncobs timely resulted in better removal efficiency – No nitrite accumulation No ammonia detected Low dissolved organic carbon concentration in treated water 10 min–3 h Nitrite and ethanol not found in treated effluent using USB reactor. With higher volumetric input rates and lesser retention times led to an increase in nitrite concentration in effluent. USB can efficiently treat water with greater hardness values 2.6–10 h Nitrosomonas europaea and Paracoccus pantotrophus were are immobilized bacterial strains used. Nitrate removal efficiency can be maintained for more than 90 days without maintenance 13 h Denitrification rate increases with increase in temperature, carbon source concentration. pH 7 is found to be ideal for denitrification 2.5 h Two-stage membrane bioreactor No nitrate and ammonium ions detected 2–17 days Soil depths have no effects on denitrification potential. Dissolved oxygen content increases with acetate addition 6h Study utilizes self immobilized microbial granules in fluidized condition. Granulation starts within 3 days from reactor start-up period

–


Bhuvanesh et al. (2013)

Fellows et al. (2011)

–

Naik and Setty (2011)

Uemoto and Morita (2010)

Green et al. (1994)

Volokita et al. (1996)

Xu et al. (2009)

Benyoucef et al. (2013)

Zhang et al. (2012)

Schipper et al. (2010a, b)

Schipper and Vukovic (2000)

Reference


250 ml

706.5 ml Polyvinyl chloride High-density polyethylene Low-density polyethylene

3l

589 ml

~1 l

Synthetic water

Synthetic water

Synthetic water

Secondary treated effluent

Municipal wastewater

HRT hydraulic retention time

8.64 l

Synthetic water

60–80

45–135

3–4.2

100

3.4–18

Starch/polycaprolactone (SPCL)

15–50

98.42

94.6

>95

100

82 >95

100

75–100

0.5–2 h

4h

3–6 h

>24 h >24 h 24 h

7h

2–6 h

2.5–6 h


Starch/polyvinyl alcohol (PVA) 7.9–8.8

Biopolymer polycaprolactone (PCL)

Polycaprolactone (PCL) Polylactic acid (PLA) and 3hydroxybutyrate-co-3hydroxyvalerate (PHBV)

Gracilaria verrucosa, polycaprolactone (PCL)

Acetate

2l

Synthetic water

Electron donor

Vol. treated

Type of water

Table 3 (continued)

Readily biodegradable COD to nitrate ratio of 2.5 realized a nitrite accumulation rate of 71.7 % Denitrification with G. verrucosa is cost effective than PCL and acetic acid. It is a complex process and requires continuous monitoring Nitrite not detected PLA is unsuitable to be used as a carbon source as no COD release during experiment Biodegradability of PHBV are much higher than PCL. PLA is difficult to degrade Sulfate ions were found to support and chloride inhibits denitrification process. Nitrite formation Polyvinyl chloride and high density polyethylene requires greater time for nitrate removal than low density polyethylene for better nitrite elimination from the effluent With prolonged operation, nitrogen removal rates declined. Highest ammonium production at boundary layer of biofilm-liquid interface. Increase in total organic carbon in effluent after long term operation Nitrite accumulation Organic carbon release Denitrification rates increased with temperature increase Initial high dissolved organic carbon and ammonia nitrogen release is observed. Diaphorobacter and Acidovorax constitute 53 % of identified denitrifying bacteria

Remarks

Shen et al. (2013)

Li et al. (2013)

Chu and Wang (2013)

Cheikh et al. (2013)

Xu et al. (2011)

Nalcaci et al. (2011)

Gong et al. (2013)

Reference



replenishment of the system with fresh source can enhance the nitrate removal efficiency (Soares and Abeliovich 1998). Greater than 95 % nitrates have been removed effectively from domestic effluent using sawdust in 5.5 days (Schipper et al. 2010a, b). Such carbon sources are very well tried to remove nutrients from drinking water (Schipper and Vukovic 2000), domestic effluent (Schipper et al. 2010a, b), and natural and synthetic groundwater (Zhang et al. 2012). Methanol and ethanol It should be noted that methanol and ethanol does not provide a solid surface for microorganism to proliferate and are not a part of solid-phase denitrification. However, methanol is one of the most common carbon sources because of its economic efficiency and availability. However, proper addition of these is compulsory; otherwise, therse can hamper effluent quality to a greater extent. It is the simplest kind of alcohol also known as methyl alcohol, wood alcohol, and carbinol (Naik and Setty 2011). All methanotrophic bacteria are strictly aerobic. They can utilize ammonia, nitrate, and sometimes nitrite as nitrogen sources (Eisentraeger et al. 2001). Methanotropic bacteria have also shown to utilize methane as both carbon and energy source under aerobic conditions. When methanol was used as the carbon source for industrial wastewater denitrification, calcium compounds used to precipitate. To avoid this precipitation, o-phosphoric acid was used along with the media to avoid this precipitation (Bhuvanesh et al. 2013). Denitrification efficiency greater than 95 % has been achieved with methanol and acetate to treat industrial wastewater in 6 h (Bhuvanesh et al. 2013). However, on testing sucrose, ethanol, and methanol, Gomez et al. in 2000 found ethanol to be the suitable electron donor source, based on experimental values. Sucrose had the lowest denitrifying rate than the above two and methanol showed toxic properties if not efficiently removed from the treated effluent.

hydroxybutyrate-co-3-hydroxyvalerate (PHBV), polyvinyl chloride, and polyethylene have also been tried as the carbon sources to the denitrifying bacteria (Nalcaci et al. 2011; Xu et al. 2011; Cheikh et al. 2013). Common disadvantages associated with heterotrophic denitrification are release of carbon sources in effluent, high bacterial growth, high biomass production, nitrite accumulation (Table 3), and cost of substrate. Organic carbon media such as wood and straw lead to denitrification with variable removal efficiencies. However, they also lead to leaching of certain unwanted substances from the media which is not taken care of such as release of phosphorous (leads to eutrophication) and greenhouse gases (CO2, CH4, N2O) (Healy et al. 2012; Tortosa et al. 2011). Healy et al. in 2012 mentioned this concept as pollution swapping where the presence of particular species goes down while the concentration of other species increases. For water and wastewater with lower BOD-tonitrogen ratio, autotrophic denitrification can be a better option than heterotrophic as external organic carbon source can be eliminated leading to reduction in cost and contamination due to lower cell yield (Zhou et al. 2011).

Solid-phase autotrophic denitrification Autotrophic denitrification is a process in which denitrifying bacteria fulfill their energy requirements like cell synthesis by reducing nitrate to elementary nitrogen gas utilizing inorganic substances as electron donors (Zhou et al. 2011). Inorganic electron donors have also found to be contributing in lithotrophic denitrification (Bottcher et al. 1990; Pauwels et al. 2000; Beller et al. 2004 and Knoller 2005). Chemolithotropic denitrification is gaining attention these days as it lowers cell production, reduces risk of biological contamination, and eliminates external carbon source and problems associated with surplus organics (Alvarez et al. 2007). Zhou et al. (2011) has divided autotrophic denitrification into three major sections:

Starch and synthetic polymers Starch is a bounteous renewable polysaccharide with easy availability, low cost, and better biodegradable properties. It is generally applied in combination with synthetic polymers such as polycaprolactone (PCL) and polyvinyl alcohol (PVA) with greater than 90 % nitrate removal efficiency (Li et al. 2013). Shen et al. (2013) reported approximately 98 % denitrification efficiency in municipal wastewater using starch in less than 2 h. Nitrate removal efficiency greater than 98 % was achieved with cornstarch and polycaprolactone blends with an average denitrification rate of 0.07 kg/m3/day (Shen et al. 2015). One of the major challenge associated with using starch as a carbon source is high COD in the treated effluent. Other polymers such as polylactic acid (PLA), 3-

(a) Hydrogen-based denitrification process (b) Anaerobic ammonium oxidation (ANAMMOX) process (c) Sulfur-based solid-phase denitrification process In a hydrogen-based denitrification process, hydrogen gas is the prime electron donor species. Among autotrophic bacteria, a hydrogen bacterium grows 24 times faster than sulfur bacteria (Zhao et al. 2011). Hydrogenotrophic denitrification has also been reported by many researchers with different reactor designs (Karanasios et al. 2010). Important disadvantages associated with the use of hydrogen as electron donor is its high operation and maintenance cost, sophisticated bioreactor setup (Chang et al. 1999; Lee and Rittmann 2002; Ho et al. 2001), and sulfate reduction (Hansen 1994). Hydrogen


gas has a low solubility of 1.6 mg/l at 20 °C in water which makes it unsafe to control and difficult to handle (Wan et al. 2009). Anaerobic ammonium oxidation (ANAMMOX) process utilizes nitrite as the electron donor species and converts ammonium to nitrogen gas (Wenk et al. 2013). Fifty percent of the nitrogen from natural marine environment is lost due to this process (Hu et al. 2011). It has been reported by Schwientek et al. (2008) that effective elimination of nitrate compounds from groundwater can be achieved by ANAM MOX and denitrification. Major disadvantages associated with this process are the slow growth rate of ammonium oxidizing and anammox bacteria (Padin et al. 2009), cost involved in maintaining anaerobic conditions, and large startup time of the reactor. Common solid-phase autotrophic denitrification sources have been discussed in detail in the following sections. Sulfur reduction Pyrite, the most abundant sulfide mineral in nature, is the potential electron donor for nitrate reduction (Schwientek et al. 2008). In a sulfur-based solid-phase denitrification process, elemental sulfur or thiosulfate acts as electron donor in presence or absence of other compounds. Many studies have reported sulfur-based solid-phase denitrification process in wastewater (Batchelor and Lawrence 1978; Driscoll and Bisogni 1978; Sahinkaya and Kilic 2014) and groundwater (Soares 2002; Alvarez et al. 2007) in combination with limestone and elemental sulfur (Flere and Zhang 1999; Zhang and Zeng 2006). Maximum water solubility of elemental sulfur reported is 0.16 μM (Steudel and Holdt 1998; Alvarez et al. 2007). Elemental sulfur is widely used for autotrophic solid-phase denitrification because of its low cost and source of alternate energy for biological denitrification (Soares 2002). Some properties of elemental sulfur make it more viable for denitrification purpose such as non-toxicity, insolubility in water, and stability in ordinary conditions (Soares 2002; Alvarez et al. 2007). Limestone, on the other hand, acts as a buffering agent and inorganic carbon source for microorganism (Soares 2 0 0 2 ; L i u a n d K o e n i g 2 0 0 2 ; Wa n e t a l . 2 0 0 9 ) . Simultaneously, it also supplements hardness in water due to the release of Ca2+ ions (Sahinkaya et al. 2011). The overall process of sulfur-based denitrification system can be demonstrated pictorially (Fig. 3). One of the major disadvantages associated with sulfurbased denitrification process is the release of sulfate along with effluent water (Sahinkaya et al. 2011). This high concentration of sulfate in treated water, if used for drinking purpose, can have laxative effects, especially in the presence of magnesium (Soares 2002). It can also cause change in the taste of water and corrosion in distribution system if it exceeds

Fig. 3 Schematic diagram of sulfur/limestone-based denitrification process. (1) Feed tank, (2) pump, (3) sulfur/limestone denitrification reactor, (4) treated effluent tank, (5) gas collection chamber

250 mg/l. USEPA guidelines suggest sulfate standard in public drinking water supply system to be 250 mg/l. Nitrite accumulation has also been detected during the process and is considered as an imprint of overdose in a sulfur-based denitrification process (Flere and Zhang 1999; Alvarez et al. 2007). Some researchers have also conducted experiments combining sulfur-based autotrophic and bioelectrochemical denitrification system (Wang et al. 2009; Wan et al. 2009). This process eliminates the usage of limestone as the hydrogen ion generated during sulfur denitrification which could be utilized in biochemical hydrogen denitrification to acquire neutralization (Wan et al. 2009). Alvarez et al. (2007) found that denitrification with thiosulphate was approximately 10 times higher than with elemental sulfur (Cardoso et al. 2006). Liu et al. in 2009 reported a combined heterotrophic and sulfurbased autotrophic denitrification system for removal of nitrate from drinking water. The process has the benefit of high efficiency from heterotrophic denitrification and sterility from sulfur-based denitrification system. Of all sulfur-based denitrifying microorganisms, Thiobacillus denitrificans is the most common (Soares 2002). Soares (2002) mentioned the sulfur-based denitrification reaction in presence of microorganisms as under 55S0 þ50NO3 ‐ þ 38H2 O þ 20CO2 þ 4NH4 þ → 4C5 H7 O2 N þ 55SO42‐ þ 25N2 þ 64Hþ

Also, the overall sulfur-based denitrification reaction as Soares (2002) and Alvarez et al. (2007) 5S0 þ 6NO3 ‐ þ 2H2 O →5SO4 2‐ þ 3N2 þ 4Hþ


A comparison of various studies for nitrate removal efficiency in sulfur-based autotrophic denitrification process at different HRTs has been shown in Table 4. Pyrite reduction Numerous studies have shown evidence of natural denitrification in conjunction with oxidation of pyrite compounds by using isotopic or geochemical measurements (Table 5) (Aravena and Robertson 1998; Cravotta 1998; Pauwels et al. 1998, 2000, 2010; Beller et al. 2004; Schwientek et al. 2008; Otero et al. 2009; Zhang et al. 2009). The process of denitrification by pyrite oxidation can be shown as (Torrento et al. 2010) 14NO3 ‐ þ 5 FeS2 þ 4Hþ →7N2 þ10SO4 2‐ þ 5Fe2þ þ 2H2 O

NO3 ‐ þ 5 Fe2þ þ 6Hþ →0:5N2 þ 5Fe3þ þ 3H2 O The overall denitrification reaction coupled with pyrite oxidation can be expressed as 15NO3 ‐ þ 5FeS2 þ 10H2 O → 7:5N2 þ10SO4 2‐ þ 5FeðOHÞ3 þ 5Hþ

Torrento et al. (2010) studied denitrification potential by varying powdered pyrite size and concentration with T. denitrificans bacteria in batch and flow-through experiments, in addition to isotopic fractionation. They found that with pyrite grain size of 25–50 μm and initial nitrate concentration of 1 mM, removal is accomplished in 14 days, whereas for higher grain size up to 100 μm and initial nitrate concentration up to 4 mM, complete consumption needed more than 60 days with release of sulfur. A release of sulfur compounds was also reported probably due to the dissolution of exterior layer of reacting species. Torrento et al. (2011) tried to study the denitrification potential of groundwater of Osona aquifer (Spain) with pyrite and T. denitrificans bacteria. They also observed that along with nitrate reduction, there was a significant sulfur release in almost all experiments, due to pyrite dissolution. They stipulated that autotrophic denitrification is prevalent mostly in batch experiments and heterotrophic in flow-through experiments. Addition of pyrite for denitrification process increased the microbial population. Torrento et al. (2011) suggested that Xanthomonadaceae can be considered as a good candidate for autotrophic denitrification and Sphingomonas, Chitinophagaceae, and Methylophilaceae for heterotrophic denitrification process in presence of pyrite as electron donor. Schwientek et al. (2008) suggested that for long-term nitrate remediation with pyrite as electron donor should not be

employed. Pyrite oxidation leads to release of iron, sulfur, and other trace metal into water that can cause aseptic conditions and finally heterotrophic infection, which can cease the overall process in the long term and hence should therefore be tested before field-scale implementation (Larsen and Postma 1997; Smolders et al. 2006, 2010; Torrento et al. 2011). Iron nanoparticle reduction Many studies have been carried out in the denitrification techniques using iron nanoparticles, with and without doping it with other metals such as nickel or palladium in groundwater and surface water (Table 5) (Cundy et al. 2008). Oxidation of iron nanoparticles produces ferrous ion and hydrogen gas which fosters the growth of microorganisms (Zhang 2003; Shin and Cha 2008; An et al. 2010; Jiang et al. 2013). At temperatures higher than 35 °C, the reactivity of Fe and Fe/ Ni nanoparticle increases and favors nitrate reduction (Ahn et al. 2008). Jiang et al. 2013 studied the impact of ironbased nanoparticles on denitrification potential of Paracoccus bacterial species under anaerobic conditions and its toxic effects on cells varying various parameters such as pH, temperature, cell growth, and other ionic species. They reported that although amalgam of Fe/Ni nanoparticle has a superior performance in nitrate reduction in presence of Paracoccus bacterial species, they also possess a higher toxicity than Fe nanoparticles only. However, the release of iron nanoparticles in water sources can pose a serious threat to both human health and environment safety (Auffan et al. 2009; Barnes et al. 2010a, b). Many soil microbes and useful bacterial communities have been reported to be toxicologically affected by iron nanoparticles (Lee et al. 2008; Diao and Yao 2009; Barnes et al. 2010a; Cullen et al. 2011; Jiang et al. 2013). The cell lysis was due to the outer coating of the cell with iron nanoparticles, reactive oxygen species generation, damage of cell membrane, and biological dechlorination (Auffan et al. 2009; Xiu et al. 2010; Barnes et al. 2010b).

Other solid-phase denitrification methodologies Bioelectrochemical process Nitrate removals from groundwater using bioelectrochemical systems have also been reported by few researchers (Clauwaert et al. 2007; Puig et al. 2011; Zhang and He 2012; Zhao et al. 2012; Xi et al. 2013; Zhang and Angelidaki 2013). The system uses electricity for the chemical deposition of nitrate ions on anode and reduces it to nitrogen gas through heterotrophic denitrification (Tong and He 2013). A heterotrophic/biofilm-electrode autotrophic solid-phase denitrification reactor was developed by Tong et al. in 2013

1.49 l

1.28 l

~0.7 l

Synthetic groundwater

Synthetic water

Groundwater

19.6 l

33.47 l

350 ml

3.2 l

Drinking water

Ground water, Beijing

Drinking water

Wastewater


Elemental sulfur, limestone

2l

Synthetic wastewater

20.9–22

30

18–102

20–700

22.6

175–750

15–60

Sulfur limestone

2.2–22.6

90

100

95

100

95.9

>98

~96

>95

~95

Remarks

Nitrate removal efficiency decreases and nitrite concentration increases with time due to biofouling. Backwashing required after certain intervals 1.5–15.4 h N2O production increased with increase in initial nitrate concentration 1h Sulfate formation to nitrate removal ratio 1.7 Sulfide not detected. Timely addition of phosphorous 0.19–6.12 h N2O production was lower. Nitrate removal rates were higher in sulfur fluidized bed than packed beds 30.5–1.8 h Denitrification rates depend on S0 surface area. Nitrite detected as intermediate product 0.3–0.6 h No accumulated nitrite and residual methanol in effluent C (methanol)/N (NO3—N) ratio 2.1–4.2 h Effluent sulfate conc. could be controlled by initial nitrate loading rate 5.6–15 h With methanol addition, Ca2+ ions in effluent declined 3h No carbon source Low sludge production High sulfate concentration in effluent HRT and temperature are key factor determining denitrification efficiency

1.5–12 h

Influent nitrate-N Removal HRT conc. (mg/l) efficiency (%)

Sulfur, limestone, methanol 50, 75

Sulfur, anthracite granules

Sulfur, limestone, activated carbon, methanol


Domestic sewage, leachate 1.5 l from sanitary landfill

Granular elemental sulfur, sodium bicarbonate

Sulfur


Vol. of water Electron donor treated

Comparison of various sulfur-based denitrification studies with their denitrifying potentials at different HRTs

Type of water

Table 4

Zhou et al. (2011)

Sahinkaya et al. (2011)

Wan et al. (2009)

Liu et al. (2009)

Alvarez et al. (2007)

Kim et al. (2004)

Soares 2002

Park et al. (2002)

Flere and Zhang (1999)

Reference


800

250

175

50

50

50

400

Synthetic water

Synthetic water

Synthetic water

Groundwater

Groundwater

Synthetic water

Synthetic water


Vol. of water treated (ml)

Zero-valent iron (ZVI)

NZVI, denitrifying bacteria

Natural pyrite crystals, denitrifying bacteria

Natural pyrite crystals, denitrifying bacteria

NZVI, hydrogenotrophic bacteria

NZVI

Zero-valent iron (ZVI), Cotton

Electron donor

40

100

24.5

38–65

50

~140

22.6, 50

Influent nitrate-N conc. (mg/l)

>94

76.36

66–100

18–52

~96

100

65–69

Removal efficiency

8 days

3 days

–

2.5 days

35 h

8–94 days

14–60 days

HRT

Comparison of denitrifying potential of past researches with ferrous or zero valent iron as the electron acceptor

Type of water

Table 5

Total organic carbon and bacterial release in the effluent Ammonium formation Nitrate and iron discharge in the effluent Nitrate removal efficiency remain unaffected in presence of high sulfate concentrations Ammonia production Ammonia production Integration of iron nanoparticles and hydrogenotrophic bacteria Nitrate reduction rate depends on pyrite grain size, pH and initial nitrate concentration. 100 % nitrate removal occurs in long term period Release of trace metals and sulfate Autotrophic and heterotrphic bacteria are stimulated through pyrite addition. Sulfate release Fe/Ni nanoparticles have higher toxicity than Fe nanoparticles. Adsorption of Fe and Fe/Ni onto cells With decrease in DO values nitrate removal efficiency increases. Transient accumulation of nitrite

Remarks

Liu et al. (2013)

Jiang et al. (2013)

Torrento et al. (2011)

Torrento et al. (2010)

An et al. (2010)

Shin and Cha (2008)

Rocca et al. (2007)

Reference



which utilizes carbon dioxide gas evolved during heterotrophic denitrification as an inorganic carbon source for autotrophic denitrification process. Zhao et al. (2011) also tried to combine heterotrophic and autotrophic denitrification through an intensified biofilm-electrode reactor. Bioelectrochemical systems have good nitrate removal efficiency for in situ groundwater remediation while it faces some other challenges like the use of purified external organic source, reduction in removal efficiency in presence of other ions, and production of brine solution. Denitrification walls Denitrification walls are porous walls dug below the groundwater table and refilled with some treating substances for groundwater contamination remediation. For the treatment of contaminated groundwater, porous treatment walls are progressively being used (Robertson and Cherry 1995; Schipper and Vukovic 2000, 2001; Hunter 2003). Construction of denitrification walls, however, did not significantly alter average nitrate content of soil in the vadose zone (Li et al. 2014). A pictorial representation of denitrification wall is shown in Fig. 4. A major limitation in groundwater denitrification is the scarcity of organic carbon source on which the denitrifying bacteria can feed. Additions of carbon sources in the sub soil layers have been implemented to increase the denitrification capacity of underground waters such as glucose and labile soil (Jahangir et al. 2012). A greater than 97 % nitrate removal efficiency has been achieved using BP-zeolite in denitrification walls (Li et al. 2014). Bates and Spalding in 1998 have tried to inject soluble carbon sources such as ethanol to shallow groundwater and evaluated the denitrification rates. Nitrite accumulation is a common problem being found while treating nitrate-contaminated groundwater with organic substrate (Table 3). Her and Huang in 1995 reported 23 % nitrite accumulation when glucose is used as a carbon source, 21 %

Fig. 4 Pictorial representation of denitrification walls for groundwater treatment

when acetic acid and 17 % when methanol is used as a carbon source for denitrification. Challenges associated with denitrification walls are as follows: unsuitable for coarse textured subsoil aquifers (Schipper et al. 2004), nitrate removal rates will vary with high and low rainfall seasons, cost of organics, and pollution swapping issue.

Discussion Nitrate remediation from groundwater is a difficult task as the sources are diffuse, control measures on different land use patterns with different rates of loading are problematic, and natural attenuation process is very slow (Tora et al. 2011). Various physical and biological conditions such as temperature, pH, DO, initial concentration, HRT, grain size, electron donor species, microbial populations, and trace and heavy metals have shown to govern the solid-phase denitrification rates in different microenvironments. Heterotrophic/autotrophic solid-phase denitrification processes are generally applied in the tertiary treatment stage of the wastewater treatment process (Zhou et al. 2011). Out of various methods of nitrate removal, solid-phase heterotrophic biological removal methods are considered to be more economical and are practiced widely (Wan et al. 2009; Jiang et al. 2013). Materials such as starch, synthetic polymers, methanol, and ethanol have been widely practiced due to their excellent denitrification efficiencies at low time periods. Moreover, it also provides simple system design and system operation. The major disadvantages linked with it are the residual carbon sources (Wan et al. 2009), cost of organics, proper dosage, high cell yield, and contamination risk. Heide et al. (2008) reported heterotrophic denitrification process as the main source of N2O generation. However, in presence of proper monitoring, regular sampling, appropriate dosage, and timely maintenance in case of starch, wood, and other low-cost natural polymers, such systems have the ability to remove most of its disadvantages and stand unmatched in terms of efficiency and cost-effectiveness. The properties of solid-phase autotrophic denitrification to produce low microbial sludge and zero secondary pollution make it a cheaper and easy to handle process than heterotrophic denitrification process. Elemental sulfur is widely used for its low cost, non-toxicity, insolubility in water, and stability in ordinary conditions. However, common problems with autotrophic denitrification like nitrite accumulation, stability in ordinary conditions, and leachate in effluent can pose toxicity that can cause threat to both human health and environment safety. Despite of shortcomings associated with every process, there are major advantages involved too, which sought to provide solution to the denitrification process. With proper doses and regular monitoring, these possible risks can be


avoided to a greater extent depending upon substrates being used. The donor species or the technologies could be adapted depending on environmental, physical, and biological conditions of the environment and must not lead to pollution swapping. However, the existence of an electron donor species with high denitrification efficiency, zero leachate, low cost, and stability in natural environment still persists and need to be focused on in the future. Acknowledgments The authors are grateful to Mr. Shashank and Miss Isha Medha for grammatical corrections and helpful feedback while framing the document.

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Neurobiology of cognitive remediation therapy for schizophrenia: a systematic review.

Effect of Chlorella sorokiniana on the biological denitrification of drinking water.

Arsenic contamination, consequences and remediation techniques: a review.

Increased nitrous oxide accumulation by bioelectrochemical denitrification under autotrophic conditions: kinetics and expression of denitrification pathway genes.

Aerobic denitrification by Pseudomonas stutzeri C3 incapable of heterotrophic nitrification.

Inhibition of existing denitrification enzyme activity by chloramphenicol.

Ion exchange membrane textile bioreactor as a new alternative for drinking water denitrification.

Stimulating in situ denitrification in an aerobic, highly permeable municipal drinking water aquifer.