Completely autotrophic nitrogen-removal over nitrite in lab-scale constructed wetlands: Evidence from a mass balance study

Chemosphere 68 (2007) 1120–1128 www.elsevier.com/locate/chemosphere

Completely autotrophic nitrogen-removal over nitrite in lab-scale constructed wetlands: Evidence from a mass balance study Guangzhi Sun b

a,*

, David Austin

b

a Department of Civil Engineering, Building 60, Monash University, Vic. 3800, Australia North American Wetland Engineering, 4444 Centerville Road, Suite 140, White Bear Lake, MN 55127, USA

Received 14 June 2006; received in revised form 23 January 2007; accepted 23 January 2007 Available online 8 March 2007

Abstract A mass-balance study was carried out to investigate the transformation of nitrogenous pollutants in vertical flow wetlands. Landfill leachate containing low BOD, but a high concentration of ammonia, was treated in four wetland columns under predominately aerobic conditions. Influent total nitrogen in the leachate consisted mainly of ammonia with less than 1% nitrate and nitrite, and negligible organic nitrogen. There was a substantial loss of total nitrogen (52%) in one column, whereas other columns exhibited zero to minor losses (<12%). Net nitrogen loss under study conditions was unexpected. Correlations between pH, nitrite and nitrate concentrations indicated the removal of nitrogen under study conditions did not follow the conventional, simplistic, chemistry of autotrophic nitrification. Through mass-balance analysis, it was found that CANON (Completely Autotrophic Nitrogen-removal Over Nitrite) was responsible for the transformation of nitrogen into gaseous form, thereby causing the loss of nitrogen mass. The results show that CANON can be native to aerobic engineered wetland systems treating wastewater that contains high ammonia and low BOD. Ó 2007 Elsevier Ltd. All rights reserved. Keywords: CANON; Deammonification; Denitrification; Nitrification; Reed bed; Wastewater

1. Introduction As a ‘green’ system with lower energy consumption, constructed wetlands have become a popular technical alternative worldwide for the treatment of various wastewaters over the past two decades (Kadlec and Knight, 1996; Brix, 1999). For example, 956 constructed wetlands have been built in the UK since the mid-1980s and recorded in a database (CWA, 2006), while approximately 200 wetlands are estimated to be operating in addition to this record. Pollutants are removed from wastewater in the wetlands by a complex variety of physiochemical and biological processes (Cooper et al., 1996; Nuttall et al., 1998). Biologically, the wetlands produce more diverse activities than *

Corresponding author. Tel.: +61 3 99055934; fax: +61 3 99054944. E-mail address: [email protected] (G. Sun).

0045-6535/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.01.060

conventional treatment systems. Although a large number of constructed wetlands have been built, very little is understood about the mechanisms for the removals of organic matter and inorganic nutrients in these systems. Current wetland designs are based on statistical analyses of performance data from existing systems, and equations (such as the Kickuth equation) which, despite having theoretical bases, in reality serve as empirical correlations of scattering data collected from the inlets and outlets of individual wetlands. Underlying statistical variability built into this design approach has led to under-sizing or over-sizing of treatment wetlands. The lack of understanding on the mechanisms of pollutant removal has also led to inadequate removal of nitrogenous pollutants from the wastewaters (Verhoeven and Meuleman, 1999; Luederitz et al., 2001). In Europe, typical percentage removal of ammonia in long-term operation is only in the range of 35–50%. To enhance nitrogen removal, it is essential that a more

G. Sun, D. Austin / Chemosphere 68 (2007) 1120–1128

comprehensive understanding is obtained on the complex nitrogen transformation processes taking place inside the ‘black box’ of constructed wetlands. Classic nitrification is a two-step process of ammonia oxidation to nitrite, followed by nitrite oxidation to nitrate under aerobic conditions (Cooper et al., 1996; Kadlec and Knight, 1996). The mass balance of this nitrification process indicates that the mass of ammoniacal-nitrogen oxidised should be equal to the mass of nitrite- and nitrate-nitrogen generated, with a small allowance for assimilatory nitrogen uptake. Classic denitrification occurs under anoxic conditions with organic carbon as electron donor and nitrate as electron acceptor. Theoretically, treatment of nitrogenous influent under aerobic conditions should result in very close influent and effluent concentrations of total nitrogen, if classic nitrification predominates and assimilatory nitrogen intake is negligible. Recent studies on the removal of ammonia have cast doubt on the traditional view that nitrogen removal in constructed wetlands primarily follows the classic pathway (Austin et al., 2003; Shipin et al., 2005). A field study by Bishay and Kadlec (2005) discovered considerable disparity between traditional nitrogen removal chemistry and the reality in a full-scale wetland system, suggesting that an alternative nitrification/denitrification route may have taken place that produced far greater ammonia removal rate than in other wetlands. In addition to the classic pathway, aerobic denitrification, heterotrophic nitrification, Anammox (anaerobic ammonium oxidation), and methane oxidation are potentially important processes in nitrogen removal in natural and engineered aquatic systems (Robertson and Kuenen, 1990; van Loodstrecht and Jetten, 1998; Strous et al., 1999; Raghoebarsing et al., 2006). The ‘new’ pathways of nitrogen transformation provide a theoretical explanation for the reason that some nitrogen mass balance studies can apparently fail despite careful execution of analytical protocols, e.g., a certain amount of nitrogen tends to ‘disappear’ when strong carbonaceous and nitrogenous wastewaters are treated (Sun et al., 2003, 2005). From another point of view, this type of failure can also provide useful information. If the analyses demonstrate imbalance of influent and effluent nitrogen for a nitrifying system, then a significant missing fraction may reveal the activity of ‘novel’ nitrogen cycle bacteria. Alternatively, a missing fraction of nitrogen may reveal abiotic processes such as ammonium adsorption (Connolly et al., 2004) or struvite (MgNH4PO4 Æ 6H2O) precipitation on media (Wang et al., 2006). Mass balance analyses force a mechanistic accounting, even if provisional, of the fate of all influent nitrogen when effluent nitrogen is not recovered by standard analytical methods. This study employs mass balance analysis to explore the transformations of nitrogen, under laboratory conditions, in vertical flow wetlands treating a landfill leachate that is characterized by low BOD and high ammonia concentrations.

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2. Materials and methods Experiments were carried out in four lab-scale vertical flow wetlands with identical dimensions, hereafter named wetland A, B, C and D. Each was comprised of a Perspex column of 95 mm in diameter and 900 mm in height and filled with 26.4 ± 7.2 mm round gravel to a depth of 150 mm as bottom layer. The main layer contained 4.4 ± 1.5 mm gravel of 650 mm depth, where a single common reed, Phragmites australis, was planted. Wetlands A, B and D were constructed in the summer of 2002 and had been used, prior to the current experiment, in studies on the treatment of agricultural wastewater. Wetland C was constructed in April 2004 and had not been used until the current experiments started in February 2005. Experiments were carried out in two periods, namely Period 1 and 2. The initial purpose of the experiments was to investigate the removal rate of ammonia from a landfill leachate. However, significant differences were found in the performances of individual wetlands during Period 1; accordingly, alterations of the experiment setup were made at the start of Period 2, and the objective of this study changed to investigate different routes of nitrogen transformation in the wetlands. In Period 1, wetlands A, B and D were arranged to form three reactors in series, whereas in Period 2 four wetlands were arranged into two parallel systems, wetlands A + B and wetlands C + D, as shown in Fig. 1. The landfill leachate was collected from a municipal waste disposal site in Portadown, Northern Ireland. In Period 1, the leachate was stored in a feed tank before being pumped into wetland A and passed through wetland B and D. In Period 2, wetlands A and C were dosed with the raw leachate simultaneously; effluents from A and C being pumped to B and D, respectively. As the risk of clogging was small due to low BOD and suspended solids concentrations in the raw leachate, all the wetlands were operated continuously without resting during the experiments. Effluent recirculation was employed individually around each wetland column. The flow rates of the feed and recirculation were carefully controlled by peristaltic pumps set to rotate at required velocities. The ratio of recirculation to the inlet flow rate was fixed at 2.5:1, while the flow rate of raw leachate was controlled at 1.35 l d1 during both experiment periods, giving a net hydraulic loading of 0.19 m3 m2 d1 on each individual wetland. At the start of Period 1, a stabilisation period of 14 d was allowed before the collection of samples began. Wastewater samples were collected from the inlet and outlet of each wetland, six sets during each period. The samples were analysed for TN, ammonia, nitrite, nitrate, COD, PO4–P, suspended solids, DO and pH. During Period 1, only three sets of samples of raw leachate and effluent from wetland D were analysed for BOD using a BODTrack apparatus (CAMLAB Ltd., UK). Because of

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G. Sun, D. Austin / Chemosphere 68 (2007) 1120–1128 Phragmite

Phragmite

Feed tank

Phragmite

Feed tank A

B

D

A

B

C

Peristaltic pump

Peristaltic pump

Peristaltic pump Period 1

D

Period 2 Vertifical flow wetlands

Vertifical flow wetlands

Fig. 1. Schematic description of the constructed wetlands in Period 1 (left) and Period 2 (right).

the small flow rate and relatively large sample volume of wastewater required, analysis of BOD was not carried out in Period 2. Ammonia was analysed using a Sension II pH/ISE meter combined with an ammonia electrode. A Piccolo pH meter was used to measure pH. DO was determined using a Sension DO electrode. The remaining parameters were analysed using a HACH DR2010 Spectrophotometer according to its standard operation procedure: TNT (test-in-tube) persulfate digestion method for TN (this method is independent of the analysis of NH4–N, NO3–N and NO2–N; persulfate digestion oxidizes all nitrogen into nitrate, which is then analyzed by a colorimetric reagent), ferrous sulfate method for NO2–N, and cadmium reduction method for NO3–N. At the end of Period 2, the roots of Phragmites in each wetland were collected and dried at 80 °C for 24 h, and dry weight was measured.

3. Results 3.1. Overall performance and dry root mass in the wetlands Throughout both periods, DO in the leachate in the wetlands was close to saturation, as shown in Table 1. Noticeably, there was little biodegradable organics in the raw leachate as can be seen in the low BOD influent concentration. Percentage removals of BOD, PO4–P and SS in different wetland columns varied considerably, from 0% to 47%, at relatively stable loadings. However, the concentrations of these pollutants were insignificant compared with ammoniacal-nitrogen (490 mg l1 in Period 1 and 483 mg l1 in Period 2), the main pollutant in the leachate. Measurement of the weight of dry roots at the end of experiment showed a much larger root mass (15.1 g) in wet-

Table 1 Mean performance and standard deviation (r) of the lab-scale wetlands Parameter

INF

Column A

Column B

Column C

EFF

D%

EFF

D%

Period 1 (n = 6) NH4–N NO2–N NO3–N TN BOD COD PO4–P SS DO pH range

490 ± 15 0.2 ± .04 1.6 ± .06 497 ± 50 23 ± 5 691 ± 69 13 ± 0.9 22 ± 2.0 7.0 ± 0.4 8.3–8.7

463 ± 7 7.5 ± 2.7 14 ± 3.0 492 ± 45

5.5 +3650 +775 1.0

434 ± 21 9.7 ± 2.7 15 ± 5.8 467 ± 29

6.2 +29 +7.1 5.1

618 ± 92 12 ± 0.6 16 ± 2 9.2 ± 0.4 8.4–8.5

11 7.7 27 +31

610 ± 71 11 ± 0.6 14 ± 4 9.7 ± 0.4 8.4–8.7

1.3 8.3 13 +5.4

Period 2 (n = 6) NH4–N NO2–N NO3–N TN COD PO4–P SS DO pH range

483 ± 22 ND 1.4 ± 0.6 485 ± 27 635 ± 81 14 ± 0.9 29 ± 2 6.5 ± 0.4 8.4–8.5

388 ± 19 42 ± 4.5 46 ± 14 477 ± 14 571 ± 47 14 ± 0.6 22 ± 4 9.5 ± 0.4 7.8–8.1

20 NA +3186 1.6 10 0.0 24 +46

328 ± 28 40 ± 8 55 ± 10 427 ± 27 530 ± 71 14 ± 1 17 ± 2 9.6 ± 0.3 7.9–8.3

15 4.8 +20 10 7.2 0.0 23 +1.1

EFF

422 ± 28 1.8 ± 1.3 1.3 ± 0.6 430 ± 50 578 ± 62 12 ± 0.5 19 ± 3 9.6 ± 0.2 7.9–8.4

Column D D%

13 NA 7.1 11 9.0 14 34 +48

EFF

D%

315 ± 25 8.7 ± 2.2 34 ± 9.1 438 ± 46 13 ± 6 552 ± 33 11 ± 1 9±3 9.9 ± 0.3 7.3–8.1

27 10 +127 6.2 43 9.5 0.0 36 +2.2

160 ± 54 2.5 ± 0.8 43 ± 14 362 ± 63 432 ± 123 10 ± 1 10 ± 3 9.7 ± 0.1 7.0–7.3

D% C 62 +39 +3208 16 25 17 47 +1.0

All values except pH are presented as mg l1. Note that BOD values are averages of three sets of data. Delta percentage changes (D%) are based on effluent from previous column (EFF) or influent (INF), whichever is the immediate preceding source.

G. Sun, D. Austin / Chemosphere 68 (2007) 1120–1128

land D, compared with in other three wetlands (A = 8.2 g, B = 7.6 g, C = 7.2 g).

3.2. The removal rate of ammoniacal-nitrogen

ing resilience to high ammonium (but not free ammonia) concentration, as in an aerated bioreactor treating high ammonia agricultural wastewater (Cannon et al., 2000). 3.3. Trend lines between pH values, NO2–N and NO3–N

On average, in Period 1 ammoniacal-nitrogen was removed by 36% (from 490 mg l1 in the raw leachate to 315 mg l1 in effluent out of wetland D) across wetlands A, B and D, which were operating in series. In period 2, NH4–N was removed by 32% (from 483 to 328 mg l1) across wetlands A–B, and 67% (from 483 to 160 mg l1) across C–D. In general, higher NH4–N removal rates (in g m2 d1, where m2 represents the surface area of a single wetland) were obtained in wetlands A, B and D in Period 2, in comparison with the rates in Period 1 (Fig. 2). This comparison could not be made for wetland C, as it was only operated in Period 2. In wetlands A and B, there was a general trend of increasing ammonia removal rate with time. In contrast, the removal rate decreased steadily with time in wetland C. In wetland D, with the highest ammonia removal efficiency, the removal rate fluctuated and appeared to peak after five weeks of operation (Fig. 3). There was no indication of nitrification inhibition in wetland D. In Period 2 the average NH4–N removal rate reached 50 g m2 d1, show-

During both periods it was found that changes in pH values of the leachate, as it flowed through each wetland column, coincided with changes in nitrite and nitrate concentrations. The decrease in pH was plotted against increase in NO2–N and NO3–N using the data collected from the inlets and outlets of individual wetlands. A straight-line correlation was established between pH and NO3–N (Fig. 4). Only a weak trend line existed between pH and the sum of NO3–N and NO2–N, whereas no clear trend was found between pH and NO2–N. 4. Discussion 4.1. Inconsistency between pH profile and the formulae of classic autotrophic nitrification It is of great interest that significant differences in ammonia removal rates are found in the lab-scale wetlands, which have the same media materials and operate under

60

-1

70 Period 1

60

Period 2

-2

-2

NH4 -N removal rate, g m d

-1

70

NH4 -N removal rate, g m d

1123

50 40 30 20 10 0

50 40 30 20 10 0

Wetland A

Wetland B

Wetland D

Wetland A

Wetland B

Wetland C

Wetland D

Fig. 2. Average ammoniacal-nitrogen removal rates during the experiment (error bars represents standard deviations).

70

70 -1

60

Wetland C

60

-2

NH4 -N removal rate, g m d

Wetland B Wetland D

-2

NH4 -N removal rate, g m d

-1

Wetland A

50 40 30 20 10 0

50 40 30 20 10 0

10

15

20

25

30

Operation time, d

35

40

45

6

7

8

9

10

11

12

13

Operation time, d

Fig. 3. Variations of ammoniacal-nitrogen removal rates with operation time (the counting of operation time begun when leachate was first introduced into the wetland).

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disparity with the implications of pH changes in the classic two-step route of autotrophic nitrification: ammonia oxidation into nitrite, followed by further oxidation of nitrite into nitrate (Metcalf and Eddy, 2003; Gray, 2004). The classic pathway consumes alkalinity at rate of 7.14 g CaCO3 for each gram of NH4–N transformed into NO3–N, whereas approximately 3.57 g CaCO3 is produced and 0.93–1.07 g organic carbon consumed in heterotrophic denitrification of 1 g NO3–N into dinitrogen gas. In the classic pathway, hydrogen ions are generated at the first step of the nitrification process, not at the second step. Therefore, if classic autotrophic nitrification is strictly followed, the concentrations of NO2–N or NO2–N + NO3–N should have a clear correlation with pH. However, Fig. 4 demonstrates that high NO2–N concentrations in the wetlands do not coincide with decrease in pH values, whereas pH appears to correlate more closely with nitrate. A previous study by Sun et al. (1998) on agricultural wastewater treatment in vertical flow wetlands showed similar result; increase in nitrate concentration (up to 2 magnitudes) appeared to coincide with decrease in pH. The pH profile in Fig. 4 indicates that the simplistic chemistry of classic nitrification (Metcalf and Eddy, 2003; Gray, 2004) is inadequate to describe complex nitrogen removal in treatment wetlands; this result being comparable to findings by Bishay and Kadlec (2005). However, the pH-N profile in Fig. 4 neither confirms nor denies alternative nitrogen pathways. More convincing evidence of the alternative pathway may include the loss of total nitrogen in the wetlands because, theoretically, nitrification via the classic route should not result in any significant loss of total nitrogen.

1.75 R 2 = 0.4829

Decrease in pH value

1.50 1.25 1.00 0.75 0.50 0.25 0.00 -0.25 -0.50 -10

0

10

20

30

40

50

60

Increase in NO3-N concentration, mg l

70

-1

1.75 R 2 = 0.0026

Decrease in pH value

1.50 1.25 1.00 0.75 0.50 0.25 0.00 -0.25 -0.50 -20

-10

0

10

20

30

40

50

60

Increase in NO2-N concentration, mg l -1 1.75 R 2 = 0.214

Decrease in pH value

1.50 1.25 1.00 0.75 0.50

4.2. CANON process indicated by nitrogen mass balance

0.25 0.00

Low BOD value (Table 1) indicates very limited biodegradable organics in the raw leachate. Negligible organic nitrogen is evident from close agreement (D 6 1%) between P the average values of persulfate TN and the sum ( N) of ammoniacal-, nitrite- and nitrate–nitrogen, P as shown in Table 2. The concentrations of TN and N in effluents from wetlands A, B, and C are also in close agreement (D 6 1.8%). The P larger difference between persulfate digestion TN and N in the effluent from wetland D may be attributable to humic substances, known to cause matrix interference for the persulfate digestion method. Treatment

-0.25 -0.50 -10

0

10

20

30

40

50

60

70

80

90

100

110

-1

Increase in NO2-N+NO3-N, mg l

Fig. 4. Correlations between decrease in pH value and concentrations of NO3–N (upper), NO2–N (middle) and NO2–N + NO3–N (lower), as the leachate passed through individual wetlands.

similar loadings. Furthermore, data analyses demonstrate that classic autotrophic nitrification pathway is not strictly followed in the removal of ammonia in these wetlands. The trend lines (Fig. 4) between pH, nitrite and nitrate show

Table 2 P Total nitrogen recovery comparison between persulfate digestion TN and N Wastewater stream analyzed

Raw leachate Effluent from Effluent from Effluent from Effluent from

wetland wetland wetland wetland

A B C D

Period 1 P N, mg l1

Persulfate TN, mg l

492 484 459 – 358

497 492 467 – 438

1

D%

Period 2 P N, mg l1

Persulfate TN, mg l1

D%

+1.0 +1.5 +1.8 – +22

484 477 422 425 205

485 477 427 430 362

+0.2 +0.0 +1.2 +1.2 +76

G. Sun, D. Austin / Chemosphere 68 (2007) 1120–1128

wetlands can produce a net increase of aromatic and oxygenated organic matter, depending on the developmental stage of the wetlands and vegetation (Barber et al., 2001). The significantly larger root mass in wetland D can produce commensurately greater amount of humic substances that exert persulfate demand, causing a falsely higher reading for persulfate TN. Thus, in the absencePof organic nitrogen (as under current study conditions), N reliably accounts for total nitrogen. P Through mass balance, the loss of total nitrogen ( N) is observed, although the degree of loss varies considerably in different wetland columns, from 1.4% to 52% (Table 3). Wetlands A and B exhibit only limited nitrogen loss; a classic nitrifying reactor mass balance results. In wetlands A and B, mass balance analysis supports nitrification as the main mechanism of NHþ 4 removal, because the mass of NH4–N removal is mostly recovered by the mass of nitrite and nitrate increase. In wetland C, where the loss of nitrogen is 12%, the removal of NHþ 4 is not accompanied by increase in nitrite or nitrate; indicating very low or zero microbial activity, which is not surprising considering the short operation history of this wetland. Initially higher, then decreasing (Fig. 3), the removal rate of NHþ 4 in wetland C is consistent with the mechanism of sorption/cation exchange moving toward saturation (Connolly et al., 2004); the sorption/cation exchange capacity of other wetlands having been exhausted due to previous operations. In wetland D, where the highest rate of nitrogen removal is achieved (Fig. 2), there is a large proportion of nitrogen ‘missing’ (ammonium removal not recovered by nitrite and nitrate), 22% Period 1 and 52% in Period 2 (Table 3). Wetland D is clearly P different from the other wetlands. It has a large missing N fraction that must be accounted for, which is provisionally named ‘a-nitrogen’ because it is uncharacterized. a-Nitrogen may be attributed to: (1) experimental error, (2) adsorption/precipitation of soluble nitrogen to solids, (3) volatilization of unionized ammonia, or (4) microbial conversion of nitrogen to biomass or gas; neither (2), (3), nor (4) could be directly detected by study methods. Experimental error does not account for a-nitrogen; in wetlands A, B and C, consistency of methods for total nitrogen anal-

Table 3 P Average loss of N as the leachate passing through individual wetlands Waste stream analyzed

Period 1 P N, mg l1

Percent loss, %

Period 2 P N, mg l1

Percent loss, %

Raw leachate Effluent out of wetland A Effluent out of wetland B Effluent out of wetland C Effluent out of wetland D

492 484

– 1.6

484 477

– 1.4

459

5.2

422

12

–

–

425

12

358

22

205

52

1125

ysis demonstrates Pprecision in measurement. Data of persulfate TN and N in Periods 1 and 2 demonstrate mass recovery of nitrogen in wetlands A, B and C, but substantial loss of total nitrogen in wetland D; the same experimental or analytical error repeated in wetland D is improbable. Adsorption in wetland D can be ruled out by comparison to wetland C. Struvite (MgNH4PO4 Æ 6H2O) formation can also be ruled out, because the required molar ratio of PO4:NH4 for the formation is 1:1, entailing very large consumption of phosphorus which is not found in Table 1. The pH in wetland D is the range of 7.1–7.8; suggesting that almost all ammonia is in ionic form and could not off-gas by volatilization unless the leachate temperature is high, i.e., above 40 °C (the temperature was not recorded, but it is believed to be below 15 °C because the experiments were carried out during winter time in the UK). Significant assimilative loss to nitrifying bacteria is improbable, as it entails yields differing by orders of magnitude in wetlands D and A for the same process. By elimination, a-nitrogen is gaseous nitrogen. High nitrate concentrations in effluent from wetland D demonstrate the occurrence of nitrification; but, what type of nitrification? The ratios of NO2–N:NO3–N are very different in the effluents out of wetlands A, B and D, 0.73, 0.91 and 0.06, respectively; suggesting that unlike in A and B, nitrite is unstable in the microbial environment in wetland D. The lack of biodegradable organics in the leachate rules out heterotrophic nitrification or denitrification. Denitrification mediated by other elements, such as sulphur, is rarely reported in wetlands. Endogenous respiration for heterotrophic denitrification would require considerable bacterial biomass to be first established then partly consumed; this has not been observed in the study system. Hence, autotrophic bacteria are most likely to be responsible for the transformation of soluble nitrogen into gaseous form. CANON (Completely Autotrophic Nitrogen-removal Over Nitrite) is consistent with the phenomena shown in wetland D. CANON is closely related to Anammox (anaerobic ammonium oxidation) process in which ammonium is first partially oxidized to nitrite, then transformed with remaining ammonium into dinitrogen by planctomycetelike bacteria (van Loodstrecht and Jetten, 1998) growing in anaerobic zones of a treatment system. Unlike the Anammox process, CANON can occur in a single-stage aerobic system, as in study conditions. Considering that wetland contains a network of intermeshing aerobic, anoxic, and anaerobic zones within the same system, it provides an ideal habitat for the coexistence of different microbial communities. The stoichiometry of the CANON process is described by Third et al. (2001) as:  þ NHþ 4 þ 0:85O2 ! 0:435N2 þ 0:13NO3 þ 1:3H2 O þ 1:4H

ð1Þ Because the CANON process generates dinitrogen gas, the loss of total nitrogen in the leachate is a natural result. The occurrence of CANON in wetland D can be further

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demonstrated if the mass of ‘missing’ nitrogen can be recovered in the mass of gaseous dinitrogen, according to stoichiometric calculation in Eq. (1). In Period 2, the mean NH4–N concentrations at inlet and outlet of wetland D are 422 mg l1 and 160 mg l1 (Table 1), respectively. Therefore, 262 mg l1 NH4–N has reacted and transformed into other forms of nitrogen. If Eq. (1) is the predominate ammonia removal pathway in wetland D, 262 mg NH4–N should produce 228 mg N2 and 34 mg NO3–N (which is close to 43 mg l1 NO3–N in the effluent measured during the experiment). Hence, the nitrogen mass that is ‘missing’ in wetland D should P be 228 mg l1 leachate treated, and the theoretical N in the effluent flowing out of wetland D should be 194 mg l1. P The measured N value in the effluent is 205 mg l1 (Table 3); thus the influent–effluent mass balance is within 5.4% when Eq. (1) is used for the mass balance calculation. Similar calculation for wetland D in Period 1 shows influent–effluent mass balance within 0.8%. These results demonstrate that, from a mass balance point of view, Eq. (1) gives an accurate description to nitrogen transformation. The ‘missing’ nitrogen is gaseous dinitrogen, and mass balance suggests that CANON is responsible for the loss of total nitrogen in wetland D. In a CANON system, aerobic (Nitroso-) and anaerobic (Planctomycete-like) bacteria can co-exist in aerobic and anoxic zones, allowing the reaction between ammonium and nitrite to occur in the anoxic zones immediately after nitrite is produced in the aerobic zones. Could such condition exist in the constructed wetlands? Although the value of DO in the leachate in this study is high, there is no reason that anoxic zones and oxygen limited conditions cannot exist inside the matrices of the wetlands. CANON can occur, and has occurred in wetland D. To understand the cause for CANON to take place significantly only in wetland D, it is necessary to look at the operation history of all four wetlands. Wetlands A, B and D were constructed in summer 2002 and had been used in studies on the treatment of high-strength agricultural effluents (Zhao et al., 2004a,b); A and B being operated under high organic loading but D under relatively lower loading. Wetland C was built in summer 2004 and had been filled only with tap water until this study began. During this study, wetland C was dosed with the landfill leachate for a short period of 12 d, which was inadequate for the establishment of a substantial microbial population. The different histories may have contributed to the differences in terms of the type and population of microorganisms in these wetlands. Third et al. (2001) reported that CANON process is subject to competition by nitrite oxidizing bacteria below ammonia loading (as N) of 0.12 kg m3 d1, whereas above that loading rate the Anammox bacteria of the CANON process out-compete nitrite oxidizing bacteria. Mean loading on each wetland in this study was 0.11 kg m3 d1, suggesting that the wetlands were operated in a transitional loading regime. In this regime, subtle differences in initial

conditions from previous operational history could be responsible for the dominance of CANON in wetland D, nitrification in A, and slight CANON activity in B (indicated by slightly higher nitrogen loss than in A). To a lesser extent, the growth of plant roots in these wetlands may also cause the differences. It is believed by many researchers that, compared with gravel beds without vegetation, the efficiency of pollutant removal is greater in vegetated wetlands (Batty and Younger, 2002; Lin et al., 2002; Karathanasis et al., 2003). In the rhizosphere of the vegetations (Macrophyte), complex interactions take place between microorganisms and pollutants, and pollutant removal could be enhanced due to more intensified microbial activities stimulated by material release via the roots. The roots can also help open up wastewater passageways to counter clogging and enhance the transfer of oxygen, which is a crucial factor in wetland performance and design (Cooper, 2005). The larger root mass in wetland D, a result of previous growth, may have benefited the microbial activities and stimulated the CANON process; however, it is not feasible to confirm and quantify the contribution by the vegetation based on the research protocol in this study. 4.3. Alternative routes of nitrogen transformations in constructed wetlands In aqueous environment, ammonia can be removed via various pathways (Robertson and Kuenen, 1990; Hippen et al., 1997; Strous et al., 1999; Jetten, 2001); the main pathway depending on a combination of factors, such as the characteristics of wastewater and availability of oxygen. As CANON is demonstrated by mass balance data in this study, other ‘novel’ removal routes (in additional to the classic pathway) may also take place in constructed wetlands. Table 4 summarises routes currently known to remove ammonia. As shown in Table 4, some processes that remove ammonia from wastewaters reduce TN concentration; these include adsorption, assimilation, heterotrophic nitrification, autotrophic denitrification, heterotrophic denitrification, and nutrient uptake by plants. It will be fascinating to investigate the potential of these pathways in constructed wetlands, which is likely to yield useful mechanistic information. It should be noted that while mass balance analysis in this study produces initial evidence to indicate the CANON process, the study is of relatively small scale and short duration. The release of dinitrogen from the wetland matrix was not monitored directly. Furthermore, the bacteria community structure of biofilm in the wetlands was not characterized. Future study needs to identify microbial community structure; this will require analytical tools traditionally used in microbiological research, such as fluorescence in situ hybridization (FISH). FISH analysis may be combined with the tracing of nitrogen flow using isotope 15 N (Kadlec et al., 2005), to yield more complete mechanistic information that chemical analysis alone cannot

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Table 4 Potential routes of ammonia transformations in constructed wetlands

Adsorption Assimilation into biomass Autotrophic nitrification Autotrophic denitrification

CANON

Heterotrophic denitrification Heterotrophic nitrification/ aerobic denitrification Methane oxidation to denitrification Nutrient uptake by plants

Brief Process Description

Contribute to TN reduction

References

Transferring ammonia from water onto the media of wetlands, typically occurring prior to nitrification Forming part of biomass generated by microorganisms during the removal of organic matter from strong wastewater Ammonia being converted into nitrite and nitrate by nitrifying bacteria under aerobic condition ANAMMOX bacteria. Under anaerobic conditions carbonate ion serves as the carbon source, nitrite as the terminal electron acceptor, to transform nitrite and ammonia into nitrogen gas In a single treatment step under aerobic but oxygen limited conditions, ammonia is first oxidized into nitrite; nitrite is then reacted with remaining ammonium into dinitrogen Nitrate and nitrite transforming into gaseous nitrogen under anaerobic condition with the presence of organic carbon Direct transformation of NH4–N to N2 or NOx species without the production of NO2–N or NO3–N. Heterotrophic nitrifiers are also aerobic denitrifiers Under anaerobic condition, prokaryotes convert methane with nitrite and nitrate into CO2 and N2 Direct uptake by plants via their roots as nutrient to be used for the synthesis of plant cells

Yes

Connolly et al. (2004) Sun et al. (2005)

provide. The analyses of a broader range of parameters (e.g., TKN and TOC) could provide further evidence, or counter-evidence, of alternative nitrogen pathways in treatment wetlands. To decrypt the complex nitrogen transformations inside the ‘black box’ of treatment wetlands an open-mind is needed. In light of recent and continuing discoveries of ‘novel’ nitrogen cycle bacteria in nature, classic descriptions of the nitrogen cycle in wastewater may not describe observed nitrogen treatment processes. The classic method of mass balance analysis, however, remain a fundamental tool for revealing which mechanisms of nitrogen treatment are relevant for a given study system. 5. Conclusions Nitrogen mass balance demonstrates that CANON took place in subsurface vertical flow wetlands treating wastewater containing high ammonia but low BOD. CANON is not an exotic process; rather, it is native to nitrifying, constructed wetlands and is likely to become a common process in wastewater engineering. Gaseous dinitrogen generated via CANON, marked as ‘missing’ in several earlier studies, is a desired end product for the degradation of nitrogenous pollutants. For nitrifying wetlands, ammonia loading (as N) greater than or equal to 0.11 kg m3 d1 can be considered to stimulate the CANON process; but the occurrence of CANON may also depend on the operating history of wetland and developmental stage of vegetation. Acknowledgement The authors acknowledge financial support from the Engineering and Physical Sciences Research Council in

Yes No Yes

Cooper et al. (1996) Strous et al. (1999); Jetten (2001)

Yes

This study

Yes

Kozub and Liehr (1999) Robertson and Kuenen (1990) Raghoebarsing et al. (2006) Prescott et al. (2002)

Yes Yes Yes

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