Nitrogen processing gradients in subsurface-flow treatment wetlands—influence of wastewater characteristics

Ecological Engineering 18 (2002) 499– 520

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Nitrogen processing gradients in subsurface-flow treatment wetlands —influence of wastewater characteristics Chris C. Tanner a,*, Robert H. Kadlec b, Max M. Gibbs a, James P.S. Sukias a, M. Long Nguyen a a

National Institute of Water and Atmospheric Research, P.O. Box 11 -115, Hamilton, New Zealand b Wetland Management Ser6ices, Chelsea, MI, USA Received 18 July 2001; received in revised form 30 December 2001; accepted 26 January 2002

Abstract Nitrification, an aerobic microbial process, is generally considered to be the rate-limiting step for N removal in subsurface-flow (SSF) constructed wetlands treating organic wastewaters. SSF wetland nitrogen (N) processing gradients were investigated using cascade mesocosms comprised of five planted (Schoenoplectus tabernaemontani ), gravel-filled tanks operated in series, in order to determine the effects of organic substrate availability (measured as chemical oxygen demand, COD) and partial pre-nitrification on rates of N removal. Duplicate sets of cascades supplied with 23 mm d − 1 of four different organic wastewaters provided COD:N ratios of  2–15 (B 1–30 in individual tanks) and a range of N species balances (34– 88% ammoniacal-N, NH4-N; B 1 – 40% oxidised N, NOx -N). Mass balances for organic N (Org-N), NH4-N and NOx -N, and COD removal were calculated for each tank of the cascades. Concurrent measurements were made of plant growth and N uptake, sediment N accumulation and selected biogeochemical indicators (redox potential, pH, and CO2 and CH4 emissions). Using a simplified model of sequential N transformations and sinks, average net rates of N mineralisation ranged from 0.22 to 0.53 g m − 2 d − 1, nitrification from 0.56 to 2.15 g m − 2 d − 1, denitrification from 0.47 to 1.99 g m − 2 d − 1 (60– 84% of measured N removal) and plant assimilation from 0.28 to 0.47 g m − 2 d − 1 in the cascade tanks. Nitrification and denitrification occurred concurrently with COD removal, even in the upstream stages of cascades receiving the higher-strength wastewaters. Surprisingly, neither net areal nitrification rates, nor first order nitrification rate constants (kA) were correlated with COD removal rates (as a measure of heterotrophic oxygen demand). Nitrification rates were correlated with average NH4-N concentrations in the cascade tanks, and were closely paralleled by net denitrification rates. Although kA for N mineralisation, nitrification and total N removal were highest for the partially pre-nitrified wastewater tested, considerably higher areal mass removals were achieved in the cascades receiving more concentrated ammonium-rich wastewaters. The oxygen demand required to support full nitrification of ammonia and mineralised Org-N in the cascades (without accounting for competitive heterotrophic oxygen demand) was in the upper range of that expected to be able to be supplied through surficial and plant-mediated oxygen transfer. Implications for understanding the nature of coupled nitrification–denitrification and COD removal in SSF treatment wetlands are discussed. © 2002 Published by Elsevier Science B.V. * Corresponding author. Tel.: + 64-7-856-1792; fax: + 64-7-856-0151. E-mail address: [email protected] (C.C. Tanner). 0925-8574/02/$ - see front matter © 2002 Published by Elsevier Science B.V. PII: S 0 9 2 5 - 8 5 7 4 ( 0 2 ) 0 0 0 1 1 - 3

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Keywords: Constructed wetlands; Wastewater treatment; Mesocosm; Nitrogen removal; Ammonification; Nitrification; Denitrification; Plant assimilation; Root-zone oxygen release

1. Introduction Extensive studies of treatment wetland performance show that areal rates of nitrogen (N) removal increase gradually with mass loading, while treatment efficiency in terms of percentage N removal gradually declines with increasing loading (Kadlec and Knight, 1996; Knight et al., 2000; Tanner et al., 1998b). Long-term average trends in loading–performance relationships are readily discernible, but there is a wide band of scatter around the central tendency, resulting in considerable uncertainty in the average final effluent N concentrations achievable at a given loading rate. Longitudinal profiles of N removal have been measured for a range of treatment wetland systems and first-order K-C* kinetic models have been used successfully to describe the observed removal gradients (Kadlec and Knight, 1996). However, the influence of different wastewater characteristics (e.g. COD:N ratios) on rate constants and background concentrations is unclear. As well as the total N concentration or load, the form of N discharged from treatment systems is also often a crucial factor affecting potential impacts in the receiving environment. In particular, ammoniacal-N (NH4-N) can be toxic to aquatic biota, and its associated nitrogenous biochemical oxygen demand (NBOD) can depress dissolved oxygen levels in waterways. However, most of the treatment wetland design models currently used are unable to accurately predict the concentrations of different forms of nitrogen discharged, or account for seasonal and stochastic variability in system response (Kadlec, 1999). Information on the relative rates and interdependency of component N removal processes occurring along the treatment gradient of wetlands is needed to improve performance predictions. The principal sustainable nitrogen removal process identified in most treatment wetland studies is

respiratory denitrification by microbes, because other potential removal mechanisms such as plant uptake, sediment adsorption and accretion, and ammonia volatilisation are generally only able to account for a fraction of the N removal rates observed (Kadlec and Knight, 1996). Respiratory denitrification involves reduction of oxidised forms of N, such as nitrate (NO3-N), to gaseous forms (predominantly N2 and N2O). However, much of the N in organic wastewaters occurs in reduced organic and ammoniacal forms, requiring initial mineralisation and/or nitrification before it can be denitrified. Subsurface-flow (SSF) treatment wetlands, in which wastewaters pass through flooded gravel beds planted with emergent species, are predominantly anaerobic environments (Bowmer, 1987). Thus, nitrification, which is an oxygen requiring process, is considered to be the primary rate-limiting step for nitrogen removal unless the wastewaters are pre-nitrified (Gersberg et al., 1983; van Oostrom and Russell, 1994). Principally, development and activity of autotrophic nitrifiers requires oxygen in addition to a supply of ammonium, alkalinity, and inorganic nutrients for bacterial growth (Kadlec and Knight, 1996). Apart from the oxygen initially dissolved within the influent wastewater, the main sources of oxygen in SSF wetlands are diffusion from the atmosphere through the water surface and, via plant shoots, into the root-zone. Rapid consumption of oxygen by heterotrophic microbes generally limits availability of this oxygen to micro-zones near the water surface and in the plant root-zone (Arth et al., 1998; Reddy and Patrick, 1984; Reddy et al., 1989). The magnitude of rootzone oxygen release and the ability of nitrifiers to access available oxygen in the organic-rich environment of a wastewater treatment wetland, where competition from faster-growing heterotrophic bacteria is likely to be fierce, is subject to question (e.g. Brix and Schierup, 1990). As well as being sensitive to temperature and pH conditions,

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nitrifier activity in a wide range of suspended and attached growth treatment systems is reported to be inhibited by high ratios of oxygen demand to reduced nitrogen, and by low (B2 g m − 3) oxygen concentrations (Metcalf and Eddy Inc, 1991). Hence, initial reduction of carbonaceous oxygen demand is often assumed to be necessary before significant nitrifier activity will be possible in treatment wetlands. We hypothesised that in SSF treatment wetlands receiving organic wastewaters: 1. Rates of nitrification and hence total nitrogen (TN) removal from ammonium-rich wastewaters would be negatively correlated with rates of chemical oxygen demand (COD) removal (as a measure of the oxygen demand resulting from degradation of organic carbon). 2. Systems receiving partially pre-nitrified wastewaters (and thus not rate-limited by this process) would exhibit higher rates of TN removal than those receiving predominantly

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NH4-N, assuming adequate organic carbon was available for denitrifiers. The objective of the present study was to test these hypotheses by comparing removal gradients of COD and key forms of N along 5-stage, SSF wetland cascade mesocosms supplied with four wastewaters with differing N to COD ratios and levels of ammoniacal- and nitrate-N. A simplified model of sequential N transformations and sinks (Fig. 1) was then used to calculate and compare net rates of plant assimilation, mineralisation (ammonification), nitrification and denitrification, and infer the oxygen fluxes required for nitrification.

2. Methods

2.1. Wetland cascade mesocosms Experiments were performed beneath a clear

Fig. 1. Nitrogen processing model schematic. The dashed arrows represent processes presumed to have little or no net effect on mass balances. Microbial biomass N, and sediment storage and sorption are presumed to be in steady state.

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Fig. 2. Diagram of a wetland cascade mesocosm comprised of five 22-l gravel-filled tanks in series. Plants have been omitted from the first tank to aid clarity. Wastewater was automatically dosed (4 d − 1) into the base of the top tank (via a vertical pipe set in the gravel), causing outflow from the surface pipe into the base of the next tank and so on down the cascade. Influents and the effluents from each stage were sampled, and the final outflow volume collected and measured, for each of the eight experimental cascades.

(  70% transmission of photosynthetically active radiation) horticultural plastic cover that excluded rainfall, at the Ruakura Research Centre, Hamilton, New Zealand (37°47%S, 175°19%E). Four pairs of cascades (Fig. 2), each comprised of five 22 l white PVC tanks in series (combined surface area 0.353 m2), were set up on a stepped platform facing northwards to minimise shading (Southern Hemisphere). Each tank had a PVC inlet pipe (32 mm I.D.) inserted to the base and a PVC surface outlet pipe (15 mm I.D.) inserted into the opposite side (30 mm from the top), and was filled to within 20 mm of the top with 15 mm diameter crushed greywacke gravel (39% porosity). This gave 8.3 l of initial free-space volume. Each tank was planted with five small vegetative propagules of soft-stem bulrush, Schoenoplectus tabernaemontani (C.C. Gmelin) Palla in January 1996. The top tank of each pair of cascades received either primary (anaerobic pond) treated meat processing wastewater (M1), primary (anaerobic pond) treated farm dairy wastewater (D1), secondary (anaerobic then facultative pond) treated

farm dairy wastewater (D2), or aerated secondary (anaerobic then mechanically aerated pond) treated farm dairy wastewater (D2A). Fresh wastewaters were collected weekly from sources nearby and transported in 200 l polypropylene barrels to the experimental site, where they were stored in a large refrigerated container (5 °C). The cascades were automatically dosed at 6-h intervals using submersible pumps, water-level controlled header tanks and solenoid release valves controlled by a programmable timer. An hour before dosing, gentle mechanical mixing of the stored wastewaters was automatically initiated to resuspend particulate material. The wastewater entered at the base of the first tank of each cascade via the inlet pipe, displacing treated water from the surface outflow pipe on the opposite side of the tank. It then flowed into the inflow tube of the next tank, and so on down the cascade. Containers placed at the end of each cascade were used to measure daily outflow volumes during the sampling period, and by difference calculate daily evapotranspiration (ET) rates.

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The cascades were supplied with 8 l d − 1 wastewater for an initial period of 10 months (January–October), whilst plants established and preliminary tests were carried out. This gave a  5 day (1 day per tank) nominal retention time for each cascade. Above-ground plant material was harvested to ground level in June (early winter) to stimulate new growth. In late October (spring), 2 weeks prior to the start of the main sampling period, 1000 l of each wastewater was collected and stored in sealed containers under refrigerated conditions (5 °C) so that a single batch of wastewater could be used throughout the sampling period. On the first day of sampling, one of each pair of cascades was also dosed at the 10 a.m. wastewater addition with 0.1 g of Br (KBr) as a conservative hydraulic tracer and 0.485 g of 15 N stable isotope as NH4Cl (15N results to be reported elsewhere).

stored in plastic bottles in an insulated container. Subsamples were analysed for pH (Orion model 9165 sure-flow combination pH electrode), ammoniacal-N (NH4-N; Orion model 95-12 ammonia electrode), nitrite N (NO2-N; Sulfanimide/NED colorimetry initiated on-site within 30 min of collection), dissolved oxidised N (NOx -N; automated cadmium reduction colorimetry), Kjeldahl N (TKN; Kjeldahl digestion, automated phenol/ hypochlorate colorimetry) and chemical oxygen demand (COD; dichromate/sulphuric acid digestion, colorimetry) using standard methods (APHA, 1989). Additional 10 ml filtered samples were stored frozen for later analysis of bromide (ion chromatography, Dionex, Sunyvale, CA). TN was calculated as the sum of TKN and NOx N, organic N (Org-N) as TKN minus NH4-N, and theoretical NBOD as 4.3 times TKN (Kadlec and Knight, 1996).

2.2. Water sampling and analysis

2.3. Gas sampling and analysis

Water exiting each tank of the cascades was sampled after the 10 a.m. wastewater addition for 5 consecutive days (to determine treatment performance) and then at 2– 4 day intervals up to day 24 to monitor the passage of tracers. Sampling wells, comprised of an 80 ml clear plastic container inserted to just above the water level in the gravel media, were positioned near the outflow of each tank and emptied by syringe prior to each sampling. When a new dose of influent was released the water level rose sequentially in each tank of the cascades, collecting a fresh sample of the out-flowing water. Redox potentials and temperature were then measured directly in the sampling wells with combination platinum and Ag/AgCl− reference electrodes (model 96-78, Orion Research, Boston, MA) and Orion model 290A meters. The redox electrodes were standardised against a ferrous– ferric standard (Light, 1972), and regularly cleaned with fine abrasive and checked against each other. Readings were taken after 2 min equilibration and converted to Eh by addition of the appropriate reference electrode temperature correction factor. Samples (60 ml) were then removed from the in situ sampling wells using a plastic syringe and

Gas emissions were sampled on the same days as water samples were taken. To reduce solar radiative heating effects, the gas space above each tank was enclosed overnight using gas-impermeable plastic bags (Tygon Vapour Barrier). These bags were sealed to the base of the tanks with a broad rubber band and suspended over a 50 cm high circular galvanised steel frame, giving a chamber volume of 0.035 m3. The bags remained sealed from 5 p.m. in the evening, until sampling at 9 a.m. the following morning (16 h), immediately prior to wastewater dosing. The outlet pipes extended to below the water level in the following tank (or the final sampling tank) to preclude gas exchange between tanks. Gas samples were withdrawn into a 60 ml syringe and transferred into duplicate 15 ml draw Vacutainers (sterile, no additive; Becton Dickinson, Franklin Lakes, NJ), allowing excess gas to vent from a second needle whilst flushing. Samples were analysed for CO2 (Mk3 infrared gas analyser, ADC Bioscientific, Hoddesdon, UK) and CH4 (Model 5890 gas chromatograph with flame-ionisation detector, Hewlett-Packard, Palo Alto, CA; separated on a 2 m, 5A molecular sieve column with a helium carrier gas flow of 25 ml min − 1). Emission rates

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were calculated assuming a linear rate of release over the enclosure time, as outlined in Livingston and Hutchinson (1995).

2.4. Plant and media sampling and analysis Plant height in each tank was maintained at or below the height of the gas enclosure frames by trimming plant shoots in the tanks to a uniform height of 40 cm whenever any rose above 50 cm in height. Trimmings were collected, dried in a forced air oven at 80 °C and weighed as a measure of plant growth. At the end of the trial, all the plant shoots were removed from each tank, counted and wet weighed. The volume of a subsample of 30 stems was determined by displacement using a 500-ml measuring cylinder (as an estimate of the chamber volume occupied, for gas emission determinations), and then dried and weighed as above. The final water-filled media volume of each tank was measured after adjusting water levels to the outlet height. Samples of the gravel media (including associated biofilm and accumulated solids) were taken from surface, middle and bottom depths (2 – 7 cm, 12–17 cm and 22– 27 cm, respectively) in the centre of each tank. The gravel was then wet sieved (2 mm mesh) to recover below-ground shoot bases, rhizomes and roots and these weighed after drying as above. Subsamples of dried above- and below-ground plant tissues, and accumulated sludge and biofilm were then analysed for N content after Kjeldahl digestion (as noted in Section 2.2).

2.5. Data analysis and process modelling Water budgets were constructed for each cascade (Table 1), partitioning measured ET losses equally between the five tanks. These were used to calculate mass balances for COD and N species in each tank of each cascade. The selected protocols for chemical and plant analysis permitted calculation of the key N transformations illustrated in Fig. 1. Nitrogen mineralisation (or ammonification) rates were calculated as net removal of Org-N plus estimated Org-N release due to turnover of below-ground plant biomass;

nitrification as N mineralisation plus net NH4-N removal less plant NH4-N uptake; and denitrification as nitrification less net NOx -N accumulation and plant NOx -N uptake. The cascade systems were relatively mature, having operated over a period of more than 10 months, and therefore sediment N accumulation and adsorption to the media and N associated with microbial biomass and particulates were assumed to be in steady state during the relatively short period of measurement. Environmental conditions in the cascades were not considered to be conducive to ammonia volatilisation (Freney et al., 1985) or dissimilatory nitrate reduction to ammonia (DNRA; Tiedje, 1988), and so these processes were assumed to be negligible. Plant uptake was calculated as above-ground harvested N plus estimated below-ground accumulation. Plant below-ground N uptake was estimated using mean below to above-ground growth ratios found for the test species in previous trials carried out with similar wastewaters under equivalent conditions (Tanner, 1994, 1996). Harvesting during the experiment removed above-ground plant tissues before they senesced, but below-ground biomass had accumulated over the 10 month establishment period. Return of N due to senescence of these tissues was estimated by multiplying the plant below-ground N mass by estimated daily turnover rates (0.00274 d − 1) derived from observations in previous studies. Plant N uptake was apportioned between ammonium and nitrate based on their relative mean concentrations in solution. Because each cascade was shown to have very near to plug flow hydraulics (see later section on hydraulics), the analyses represented the spatial information along a gradient from inlet to outlet, at intervals of one-fifth of the full length. Mass loadings and removals for each nitrogen species for each tank, together with concentrations along the flow path from tank to tank, were combined to calculate the first order areal rate constants for the individual processes along the gradient from inlet to outlet, considering each tank to be a differential reactor (Levenspiel, 1972). The rate constant (m y − 1) was computed to be the process rate (g m − 2 y − 1) divided by the mean con-

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centration in the tank (g m − 3), utilising the data for the various nitrogen species in the water along each of the cascades. The K-C* model (Kadlec and Knight, 1996) was fit to data for organic, ammonium and TN. Nitrate was present in insufficient concentrations to permit profile fitting, but mass balances allowed calculation of the effective rate constant in each tank of each cascade.

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3. Results and discussion

3.1. Influent wastewater characteristics The overall areal mass TN loadings applied to the experimental cascades (Table 2) were well within the range commonly applied to SSF treatment wetlands (Kadlec and Knight, 1996). The differing sources and levels of preceding treatment in the wastewaters applied to the cascades pro-

Table 1 Hydraulic characteristics of the experimental cascades Cascade stage

M1 1 2 3 4 5 O6erall D1 1 2 3 4 5 O6erall D2 1 2 3 4 5 O6erall D2A 1 2 3 4 5 O6erall

Inlet flow (l d−1)

Outlet flow (l d−1)

OpenVoids (%)

Root-filled voids (%)

Sediment-filled voids (%)

HRT (d)

HLR (m d−1)

8.00 7.55 7.11 6.66 6.21 ET 6.3 mm d−1

7.55 7.11 6.66 6.21 5.76

54 48 55 61 61 56

22 17 25 25 32 24

24 35 20 15 8 20

0.58 0.54 0.67 0.78 0.84 3.41

0.113 0.107 0.101 0.094 0.088 0.023

8.00 7.67 7.34 7.01 6.68 ET 4.7 mm d−1

7.67 7.34 7.01 6.68 6.35

37 36 51 70 68 53

15 25 21 22 21 21

48 39 28 8 11 27

0.40 0.40 0.59 0.85 0.87 3.12

0.113 0.109 0.104 0.099 0.095 0.023

8.00 7.38 6.77 6.15 5.53 ET 8.7 mm d−1

7.38 6.77 6.15 5.53 4.91

62 66 71 78 77 71

35 21 20 20 19 23

3 13 9 2 5 6

0.67 0.77 0.92 1.10 1.22 4.68

0.113 0.104 0.096 0.087 0.078 0.023

8.00 7.45 6.90 6.35 5.81 ET 7.8 mm d−1

7.45 6.90 6.35 5.81 5.26

67 65 67 75 76 70

28 16 12 12 10 16

5 18 20 12 14 14

0.72 0.76 0.84 1.03 1.14 4.49

0.113 0.105 0.098 0.090 0.082 0.023

Each tank contained 20.5 l of gravel with an initial void fraction of 8 l (39%). Each cascade received 8 l d−1 of wastewater giving an initial theoretical hydraulic retention time (HRT) of 1 day per tank and 5 days overall for the cascade.HLR, hydraulic loading rate; ET, mean evapotranspiration loss recorded during the experimental period.

1.80 1.39 1.28 0.97 0.30 1.15 93%

1.56 1.67 0.45 0.18 0.05 0.78 81%

0.29 2.26 3.66 3.25 2.42 2.38 76%

2.25 2.00 1.56 2.36 3.24 2.28 50%

TN removal

0.34 0.33 0.38 0.22 0.04 0.26

0.12 0.14 0.15 0.09 0.06 0.11

0.02 0.02 0.06 0.12 0.19 0.08

0.00 0.00 0.01 0.02 0.08 0.02

NOx -N uptake

0.25 0.11 0.08 0.16 0.11 0.14

0.63 0.59 0.31 0.20 0.11 0.37

0.52 0.49 0.69 0.50 0.34 0.51

0.45 0.52 0.78 0.75 0.83 0.66

NH4-N uptake

0.58 0.44 0.46 0.38 0.16 0.40

0.75 0.73 0.46 0.29 0.17 0.48

0.55 0.51 0.75 0.62 0.53 0.59

0.45 0.53 0.78 0.77 0.91 0.69

TN uptake

Plant uptake and turnover

0.19 0.17 0.12 0.11 0.03 0.12

0.33 0.23 0.14 0.08 0.05 0.16

0.25 0.18 0.27 0.15 0.16 0.20

0.19 0.17 0.25 0.24 0.25 0.22

Belowground N turnover

0.39 0.27 0.33 0.27 0.13 0.28

0.43 0.50 0.32 0.21 0.12 0.31

0.29 0.32 0.48 0.47 0.37 0.39

0.26 0.35 0.54 0.53 0.67 0.47

Net TN uptake

0.63 0.44 0.47 0.31 0.12 0.39

1.08 0.51 0.23 0.24 0.07 0.43

−1.33 0.53 1.08 0.47 0.36 0.22

0.64 0.44 0.48 0.50 0.59 0.53

Net mineralisation

1.10 1.16 0.74 0.17 0.04 0.64

1.39 1.11 0.25 0.06 0 0.56

0.26 1.75 3.49 3.13 2.13 2.15

2.03 1.68 1.05 2.08 3.15 2.00

Net nitrification

Microbial processes

1.41 1.12 0.94 0.69 0.17 0.87

1.14 1.17 0.13 0 0 0.47

0 1.93 3.17 2.78 2.05 1.99

1.99 1.65 1.02 1.84 2.57 1.81

Net denitrification

5.04 5.29 3.37 0.77 0.19 2.93

6.35 5.05 1.15 0.27 0 2.55

1.18 8.02 15.94 14.31 9.73 9.84

9.26 7.67 4.78 9.49 14.41 9.12

Consumed by nitrification (NBOD)

4.02 3.20 2.70 1.98 0.49 2.48

3.24 3.35 0.37 0 0 1.39

0 5.52 9.07 7.95 5.86 5.68

5.69 4.71 2.93 5.24 7.35 5.18

1.02 2.10 0.67 0 0 0.76

3.11 1.70 0.77 0.27 0 1.17

1.18 2.50 6.87 6.36 3.87 4.16

3.58 2.96 1.85 4.24 7.06 3.94

Equivalents Net NOD returned in for coupled denitrification process

Oxygen demand

All rates are g m−2 d−1 unless otherwise indicated. Small indeterminate values replaced by zero. a Mean mass TN loading and percentage removal for complete cascade; i.e. initial loading spread across all 5 stages of each cascade and removal cumulative along cascades.

6.2 4.4 3.0 1.7 0.7 3.2 1.2

D2A 1 2 3 4 5 Mean Overalla

15.6 15.3 13.0 9.4 6.1 11.9 3.1

D1 1 2 3 4 5 Mean Overalla

4.8 3.2 1.5 1.1 0.9 2.3 0.96

22.6 20.4 18.4 16.8 14.4 18.5 4.5

M1 1 2 3 4 5 Mean Overalla

D2 1 2 3 4 5 Mean Overalla

TN loading

Cascade stage

Table 2 Summary of nitrogen mean areal mass removal and process rates for four wastewaters in the experimental cascades

78 80 74 72 57 – 76

73 70 29 0 0 – 60

0 86 87 86 85 – 84

89 82 66 78 79 – 79

Denitrification (%)

22 20 26 28 43 – 24

27 30 71 100+ 100+ – 40

100 14 13 14 15 – 16

11 18 34 22 21 – 21

Net plant uptake (%)

Relative TN losses

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Fig. 3. Comparison of mean (a) nitrogen species and (b) oxygen demand characteristics of the four experimental wastewaters.

vided a range of organic matter to N ratios and N forms (Fig. 3). The primary meat processing and dairy wastewaters (M1 and D1) both showed high levels of TN with a marked predominance of NH4-N, but differing relative COD levels (low and high, respectively). The secondary and secondary aerated dairy wastewaters (D2 and D2A) showed roughly similar levels of TN and COD, but differing ratios of N forms. In contrast to the other wastewaters, D2A contained substantial levels of NOx -N (40% of TN). The theoretical additional oxygen requirement (NBOD) for complete nitrification of NH4-N and potentially mineralisable Org-N was particularly high for M1, comprising nearly 70% of the total oxygen demand (Fig. 3b). In comparison, the NBOD of the dairy wastewaters (D1, D2 and D2A) comprised  20 – 35% of their total oxygen demand.

3.2. Hydraulics Three features of the experimental systems led to near plug flow behaviour: (1) overflow displacement from tank to tank, (2) intermittent dosing, and (3) the vertical upflow mode. Br tracer testing showed that a steady-state 10 tanksin-series model gave a good fit to the experimental data (compare D2 results with theoretical, Fig. 4).

The tracer tests showed close to mass balance closure (e.g. 100 mg added versus 115 mg collected for D2), as well as close-to-theoretical detention time for the D2 and D2A cascades that accumulated minimal organic solids (Table 1). The initial two tanks receiving the higher strength M1 and D1 wastewaters had more accumulated sediments, sludge and/or roots than the downstream tanks, and free water volumes in these tanks were reduced by 46–64% (Table 1). Evidence of this reduced volume and some shortcircuiting was obvious in the Br tracer response of these initial tanks (illustrated for D1, Fig. 4). In contrast, free water volumes in the first D2 and D2A tanks were only reduced by 33–38%. This root and sediment accumulation in the gravel media, along with cumulative ET losses resulted in individual tank detention times increasing with distance through each cascade (Table 1). Overall detention times for the cascades ranged from 3.1 to 5.5 d at the hydraulic loading rate of 0.023 m d − 1 applied to all cascades.

3.3. Biogeochemical indicators Nitrification of NH4-N results in consumption of alkalinity, whilst denitrification generates alkalinity (Reddy and Patrick, 1984). The influence of

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these processes on pH depends on their relative rates and the buffering capacity of the wastewater. Concurrent decomposition processes (e.g. organic acid generation from anaerobic fermentation) and H+ release from plant roots assimilating NH4-N (Marschner, 1995) will also interact with these processes. M1 and D1 showed small reductions in the pH of outflowing surfacewaters during passage along the cascades, but pH remained above 6.9 (Fig. 5). In the lower-strength D2 and D2A wastewaters pH declined to 6.3 and 6.7, respectively, in the second and third

Fig. 4. Tracer response through D1 and D2 cascades, compared to theoretical results for ten cells in series (two per actual tank). Results for the D2 cascade shows close to theoretical tracer response, within the limitations of sampling frequency. The tracer response for the D1 cascade shows considerable lost volume due to clogging with organic matter and root growth, as evidenced by faster tracer transit.

tanks of the cascades, then rose slightly in the later tanks. Redox potentials, which were only measured in the surface-waters of the cascade tanks, showed relatively oxidised conditions (Eh\ 300 mV) for all the dairy wastewaters, with slightly lower values for D1 than D2 and D2A (Fig. 5). Redox conditions in the initial tanks receiving M1 wastewater were substantially more reduced (mean Eh 0 mV). This coincided with markedly higher methane emissions from M1 (see below) and the presence of black sulphurous colloidal matter, in the outflows of the first two tanks of these cascades. In previous studies with these wastewaters, total sulphur levels of B 10 g m − 3 have been measured for D1 (Tanner, 1996), compared to 50 g m − 3 or more for M1 (van Oostrom, pers. com.; due to high S content of animal proteins in this waste and use of sulphuric acid for fat rendering). Markedly lower redox potentials are likely to have occurred deeper in the cascade tanks (Tanner et al., 1997). Overnight carbon dioxide and methane emissions were measured as an indication of organic matter processing rates along the cascades (Fig. 6). Measured CO2 emissions showed large day-today variation, with highest median levels in the M1 and D1 cascades. Gas transport through S. tabernaemontani occurs predominantly by simple diffusion through aerenchymous tissues, rather than convective flow, and is thus not likely to have been greatly affected by diel environmental cycles (Brix et al., 1992; Tanner et al., 1997). However, variations in CO2 emission rates were found to be negatively correlated (R 2 = 0.65–0.88 for individual cascades) to the mean morning solar radiation level recorded at the nearby ( 100 m) meteorological station. This suggests that, particularly on sunny mornings, plant photosynthetic uptake was markedly depleting overnight CO2 accumulations, and therefore the measured overnight CO2 emission rates are only useful as qualitative indicators of organic matter processing activity in the tanks. Carbon emissions as methane were generally around an order of magnitude less than those recorded for CO2 (Fig. 6). They were markedly higher in the initial two tanks of the M1 cascades

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Fig. 5. Gradients of (a) redox potential and (b) pH along the four experimental cascades. Values are means over 5 days for 2 duplicate cascades ( 9standard deviation) sampled at the inlet (stage 0) and the outlet (stages 1 – 5) of each tank.

(median 30 and 25 mgC m − 2 d − 1, respectively), which showed the lowest redox potentials. CH4 emissions declined to levels less than one third of this in later stages of these cascades. Median CH4 emissions declined from 8 to 3 mgC m − 2 d − 1 along the D1 cascades, but remained at similar levels of  4.5 mgC m − 2 d − 1 along both the D2 and D2A cascades. These latter rates were similar to median mid-summer methane emissions of 2– 6.4 mgC m − 2 d − 1 recorded in pilot-scale SSF wetlands planted with the same species treating the D2 wastewater (Tanner et al., 1997). Although methanogens are generally considered to be strict anaerobes requiring highly reduced conditions (EhB −300 mV) for growth, co-existence with aerobic processes is commonly observed in many natural sediments and wastewater treatment systems (Zitomer, 1998). Plant root-zone oxygen release was likely to have suppressed the development of redox conditions conducive to methanogenesis and enhanced the potential for CH4 oxidation in these systems (Tanner et al., 1997).

3.4. Plant growth and N assimilation Plant growth was vigorous in all tanks, with above-ground growth ranging from 8 to 24 g m − 2 d − 1 (12–44 g m − 2 d − 1 including estimated below-ground production) during the study period. Harvesting of shoots at 40 cm above the gravel surface appeared to have no detrimental effects on above-ground growth. Plants kept growing from basal meristems and there was no sign of chlorosis or necrosis of cut shoots. Roots penetrated to near the base of the tanks (300 mm), except in the initial tanks receiving the higher strength M1 and D1 wastewaters, where there was some blackening and stunting of roots. There was also evidence of lower plant biomass in these tanks. This was unlikely to be due to ammonia toxicity, because of the near neutral pH of the wastewaters. Other properties (see later), such as hypoxic stress at low redox potential (Kludze and DeLaune, 1995) or sulphide toxicity (see later, Koch and Mendelssohn, 1989), may have inhibited growth.

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Plant N concentrations showed a general rise with average wastewater N concentration, ranging from 11 to 30 g kg − 1 in above-ground tissues and 7–29 g kg − 1 in below-ground tissues. These covered nearly the complete range reported for S. tabernaemontani in treatment wetlands (Tanner, 2001a), suggesting plants experienced a broad range of N availability along the cascades. This species has been shown to be able to utilise both NH4-N and NOx -N (Hally, 1990; Tanner, 1994). With ammonium:nitrate ratios in the tanks of the cascades varying from 0.2 to 252, nitrate was likely at times to be an important or dominant source of N for plant growth (Table 2). Net plant N uptake (uptake minus belowground turnover losses) was generally a minor component of TN removal in the M1, D1 and D2A cascades representing less than 25% of over-

all removal (Table 2). It became relatively more important representing all of the TN removal in the initial tank of D1, where overall TN removal was low, and in the final tanks of the D2 cascades, where dissolved inorganic N had been reduced to low levels and remaining N was predominantly in slowly mineralising organic forms. Plant growth in the experimental tanks was likely to have been stimulated by shoot-tip harvesting and ‘edge-effects’, particularly side-lighting (e.g. Tanner, 1996). Accordingly, average net plant N uptake rates of 0.48 gN m − 2 d − 1 recorded for the D2 cascades were relatively high compared to values reported for this species in larger-scale established wetlands treating the same wastewater (e.g. 0.2–0.3 gN m − 2 d − 1; Tanner, 2001a). However, presumably because the plants

Fig. 6. Box and whisker plots of CO2 and CH4 emission rates for the cascades over a 5 day period. The boxes delineate the interquartile range, above and below the median (central horizontal line), and the ‘whiskers’ show the overall range of the data.

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Fig. 7. Comparison of COD and TN concentration gradients along the wetland cascades. Other details as for Fig. 5.

had well established rhizome systems which had largely filled the available growing space, N uptake rates were only about one third of those reported for this species in experimental systems where plants were still actively spreading (Tanner, 1996). This suggests that plant N uptake rates in the cascades were moderately, but not excessively, above normal levels likely to occur in treatment wetlands.

3.5. COD and nitrogen attenuation Profiles of COD and TN concentration along the cascades are compared in Fig. 7. D1 showed concentration reductions of over 1000 g m − 3 COD and 100 g m − 3 TN down the cascades. Smaller reductions in COD ( 100 – 250 g m − 3) and TN (30 –60 g m − 3) concentration were recorded for the other wastewaters. During the monitoring period M1 and D2 showed 10 – 20% increases in COD (Fig. 7) and D1 50% increase in Org-N during passage through the first tank of the cascades (Fig. 8). Apart from this latter case,

Org-N and NH4-N showed gradual reductions down the cascades (Fig. 8). Relatively low or negative rates of net COD removal in the initial tanks of all the cascades are likely to be due to regeneration of soluble or colloidal COD from accumulated organic solids (Table 1). Similarly, increase in Org-N concentrations in the initial D1 tanks is likely to have resulted from mineralisation of the large organic matter accumulations in these systems. NOx -N present in relatively high concentrations in D2A declined markedly from  22 to 1 g m − 3 down the cascades (Fig. 8). Apart from a small increase after passage through the first stage, NOx -N concentrations remained low along the D2 cascades. In contrast, NOx -N rose gradually in the later stages of M1 and D1, reaching around 8–10 g m − 3 at the outflow. Alkalinity levels were adequate (pH remaining above 6.8; Fig. 5b), but associated levels of COD removal were low, suggesting denitrification was likely to be limited by organic carbon availability in the later stages of these cascades.

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Greatest TN and COD mass removal was recorded for the D1 systems, peaking in the middle tanks of the cascades at 3.6 and 66 g m − 2 d − l, respectively (Fig. 9). M1 systems showed intermediate rates of TN removal, but, in contrast, TN removal was lowest in the middle of these cascades. The D2 and D2A cascades showed similar patterns of TN removal despite differences in wastewater N speciation and COD concentration. TN mass removal showed a general decline along these cascades.

3.6. Comparison with pilot-scale wetland performance The D2 wastewater was the subject of a 5-year study carried out  3 km away (Tanner et al., 1998b). Five pilot-scale gravel beds (19 m2) planted with the same species (S. tabernaemontani ), were operated at five different hydraulic loading rates of D2 wastewater. Comparison with the performance in the D2 and D2A cascades in present study better TN removal (reduced outflow concentrations in relation to TN mass loading) for the experimental cascades (Fig. 10). Treatment of similar M1 wastewater was previously studied over a 2 year period in pilot-scale (18 m2) SSF wetlands 13 km from the experimental site (van Oostrom and Cooper, 1990). Annual COD removal recorded during this study

in gravel beds (18 m2) planted with the same species was 5.3 g m − 2 d − l at loading rates of 7.6 g m − 2 d − l (70% removal). This was slightly less than the mean removal rate of 6.5 g m − 2 d − l at loading rates of 8.7 g m − 2 d − l (76%) recorded for the M1 cascades in the present study. In December of the second year of operation, TN removal in the pilot-scale wetlands was 2.1 g m − 2 d − 1 at a mass loading of 4.1 g m − 2 d − 1 (51%, Hally, 1990). This is comparable to the TN removal in the M1 cascades during the present study (1 month earlier in the season) of 2.6 g m − 2 d − 1 at a mass loading of 4.5 g m − 2 d − 1 (58% removal). The annual TN removal performance reported by van Oostrom and Cooper (1990) for these pilotscale wetlands was, however, lower (0.8 gN m − 2 d − 1; 20%) than found in the present study. This may be partly due to residual effects from previous operation at considerably higher TN loadings (8.1 gN m − 2 d − 1) in that study. The above comparisons suggest that the cascade mesocosms provided slightly better COD and moderately better TN removal than for larger pilot-scale horizontal SSF systems. This is likely to be mainly due to the improved hydraulics in the multi-stage cascades. The design of the system promoted vertical water movement through each stage of the cascades (see section 3.2). This provided hydraulic conditions closer to plug-flow than that commonly achieved in full-scale SSF

Fig. 8. Concentration gradients of constituent N species along the cascades. Other details as for Fig. 5.

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Fig. 9. Gradients of key nitrogen transformation and plant uptake rates compared to TN and COD removal along the four experimental cascades. Note: separate scale applying to COD removal.

treatment wetlands (e.g. Bowmer, 1987), and forced wastewaters to flow through surficial zones of the media with higher density of plant roots and greater potential oxygen availability. Tracer studies in the pilot-scale D2 wetlands (Tanner et al., 1998a) showed considerable longitudinal dispersion (indicating non-plug flow), with clogging of the media causing markedly reduced mean residence times in the higher-loaded wetlands.

The performance of the cascade systems may have also been affected by ‘beneficial’ edge effects resulting from factors such as: increased levels of lateral lighting (promoting increased plant growth and nutrient uptake; see Section 3.4)) and advection (elevating evapotranspiration via ‘clothesline’ and ‘oasis’ effects), and ramification of rhizomes against tank walls. For these reasons, our experimental results should not be indiscriminately extrapolated to larger-scale systems.

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Fig. 10. Comparison of mean TN performance data for D2 and D2A cascade tanks in the present study with that for previous pilot-scale wetlands treating similar D2 wastewaters (Tanner et al., 1998b). The trendlines are linear regressions (R 2 = 0.96 Pilot, 0.89 Cascades).

3.7. Nitrogen processing gradients N processing rates derived from our model showed markedly different patterns along cascades treating different wastewaters (Fig. 9). Contrary to our initial hypothesis, neither net areal nitrification rates, nor first-order nitrification rate constants showed clear correlations with corresponding COD removal rates (Fig. 11a) or concentrations. Because of differences in the degradability of COD in different wastewaters, COD removal rates (: oxygen demand exerted) are likely to better reflect the availability of or-

ganic substrates and thus potential heterotrophic competition for oxygen. In accordance with our second hypothesis, calculated first order areal rate constants (kA) were highest for net organic, ammoniacal and total N removal in the D2A cascades treating partially nitrified wastewaters (Table 3). Org-N (and thus TN) background concentrations (C*) were also lowest in D2A. Although these cascades, which received the lowest COD loadings, showed a  80% increase in areal denitrification rates over the D2 cascades receiving similar un-nitrified wastewaters (0.8790.41 cf. 0.479 0.59 gTN m − 2 d − 1), their mean TN removal efficiency in relation to loading was similar at  65% for both D2 and D2A. The highest areal TN mass removals and denitrification rates were recorded for the cascades receiving the higher strength, un-nitrified M1 (2.390.71 gTN m − 2 d − 1) and D1 (2.49 0.67 gTN m − 2 d − 1) wastewaters. Thus, although pre-nitrification increased the areal efficiency of nitrogen removal in relation to TN concentration (kA), it did not substantially increase areal mass N removal rates (g m − 2 d − 1). This is contrary to the concept that TN removal from ammonium-rich wastewaters is controlled, via nitrification rates, by a (relatively limited) potential maximum areal oxygen flux (via direct and plant-mediated diffusion) into the wetland media. Modelled net nitrification rate constants (kA) and net areal nitrification rates both showed an initial increase followed by a gradual decline with rising average NH4-N concentrations (Fig. 11b).

Table 3 Mean first order, areal N processing rate constants (kA) and background concentrations (C*) in the experimental cascades (9 standard deviation) based on modelled net N transformations Wastewater

M1 D1 D2 D2A Literature rangea

Net N mineralisation

Net nitrification

Net denitrification

Total N removal

kA (m y−1)

C* (g m−3)

kA (m y−1)

kA (m y−1)

kA (m y−1)

C* (g m−3)

15.0 9 4.4 11.2 9 5.0 12.9 97.6 21 9 6 20–57

4.0 5.3 2.2 0.7 0–5

4.8 9 2.5 19 912 18 9 15 35924 0–33

380 9 273 142 9 103 133 9 99 3096 6–215

5.4 91.9 11.8 97.5 14.1 9 11.2 16.1 9 4.5 5–37

4.0 5.3 2.2 0.7 0–10

C* set to zero for nitrification and denitrification. a Literature values from Kadlec and Knight (1996) for SSF wetlands, and Knight et al. (2000).

Fig. 11. Relationship of net areal nitrification rates and nitrification rate constants (kA) to (a) COD areal mass removal (as a measure of heterotrophic oxygen demand) and (b) average NH4-N concentration in the cascade tanks. The smoothed trendlines shown in (b) were fitted using a locally weighted (50%) least squares method.

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The kA for nitrification peaked at average NH4-N concentrations of 20 – 40 g m − 3 and areal mass rates at 80–140 g m − 3. The apparent decline in net areal nitrification rates at high NH4-N may have been due to inhibition by other factors associated with the ammonium-rich M1 wastewaters, such as elevated sulphide levels (as previously discussed in relation to plant growth). The presence of black sulphurous colloidal material in the water leaving the two initial M1 tanks, and the substantially lower surface-water redox potentials and higher methane emissions shown by these tanks, suggest that both sulphate reduction and methanogenesis were important degradation processes in these systems. These, generally competitive, processes can occur together (Capone and Kiene, 1988) where levels of sulphate (the thermodynamically superior electron accepter) are low relative to available electron donors (e.g. organic matter). Overall measured areal mass N removal in the cascades in relation to mass loading in the present study (Table 2) were in the mid to high range (Tanner, 2001b), but areal rate constants low to moderate (Table 3), compared to rates reported for other planted SSF wetlands. Inferred rate constants for N mineralisation in the cascades were generally low compared to results reported for other SSF treatment systems, while those for nitrification and denitrification covered much of the reported range (Table 3). Rate constants were notably high for nitrification in D2A and denitrification in M1.

3.8. Validity of key model assumptions The model ignored volatilisation of ionised ammonia and DNRA, because available evidence suggested these were unlikely to be important mechanisms of N loss or transformation, respectively, in the cascade wetlands. Volatilisation rates estimated (Freney et al., 1985) for the cascades using average NH4-N concentration in the tanks, measured surface pH and temperatures, and hourly wind records from a nearby (100 m) meteorological station were very low (M1 and D1B 0.041; D2 and D2AB 0.006 g m − 2 d − 1). Furthermore, measurements

made in similar pilot-scale wetland treatment systems treating M1 wastewater (which in the present study showed the highest surface-water pH and NH4-N concentration, and thus was the most likely to experience volatilisation) found volatilisation to be negligible (maximum 0.0048, mean 0.0014 gN m − 2 d − 1; Hally, 1990). DNRA is generally considered to be favoured in low redox, electron accepter-poor and organic carbon-rich environments (Tiedje, 1988). van Oostrom and Russell (1994) found DNRA to represent less than 5% of NOx -N removal in experimental constructed wetland systems treating nitrate-rich meat processing wastewaters. Bowden (1986) and Nijburg and Laanbroek (1997) reported DNRA accounting for B10% of nitrate losses in freshwater wetlands. Any DNRA occurring in the cascades would have elevated the rates of nitrification (and concomitant oxygen supply) required to account for measured N losses in the cascade wetlands. A potential weakness in our model was our presumption that the sediment and biofilm components of the wetland were stable, with no N net accumulation or losses during the monitoring period. Having operated over an extended period with stable flows, it is likely that sorption, desorption would have been in equilibrium with the wastewater in the tanks, and immobilisation in microbial biomass relatively stable. However, a substantial quantity of particulates had accumulated, particularly in the upstream D1 and M1 tanks (Table 1), and mass balances (Fig. 9) showed net generation (negative removal) of COD (except D1) and Org-N (only D1) in these tanks. TN removal via sedimentation, which was estimated (based on preliminary measurements of SS removal in the cascades) to account for B3.7% of measured TN removal in the M1 and D1 tanks and B 1.8% of measured TN removal in the D2 and D2A tanks, was not considered to be an important removal mechanism. However, TN accumulated in the media at the end of the study (after 300 days operation) ranged from 20 to 74 g m − 2 in the M1 and D1 tanks and 6–55 g m − 2 in the D2 and D2A tanks (both generally highest in upstream tanks). This was far in excess (5–130fold greater) of the N accumulation predicted

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from sedimentation alone, except in the initial tanks where it was similar to predicted cumulative sedimentation. Such elevated N accumulation in the media of the tanks suggests considerable immobilisation in biofilms and/or contribution from below-ground plant litter. It also, particularly in the higher-loaded and upstream tanks, suggests the potential for considerable regeneration of accumulated organic matter and N. Our model only accounted for net loss of COD and TN measurable within the respective tank outflows and that estimated from recycling of below-ground plant biomass N. If net regeneration from accumulated (non-plant) organic matter in the media also occurred (as seems likely, at least, in the upstream tanks), this would require even higher rates of nitrification and denitrification to account for the observed net N mass balances. S. tabernaemontani has been shown to utilise both ammoniacal and nitrate nitrogen (Hally, 1990; Tanner, 1994), and nitrate reductase activity is generally rapidly inducible in plants (Marschner, 1995). Ammonium:nitrate ratios in tank waters varied from 2.2:10.8 to 177:0.7, and thus both forms of nitrogen were likely at times to be of importance in plant uptake (Table 2). In the absence of specific information of the relative uptake kinetics of these N species by S. tabernaemontani, we took the simple approach of apportioning plant N uptake between these two forms based on their relative mean concentrations in each tank. Although the situation is likely to be more complex than this (Marschner, 1995), plant uptake was only a major mechanism of N uptake in the downstream stages of the D2 and D2A cascades and it is only here that differences in relative uptake would have more than a minor effect on rates of other N removal processes.

3.9. Oxygen demand implications of obser6ed nitrification rates If the overall assumptions of our model are reasonable, then the nitrogenous oxygen demand exerted (4.57 g O2 g − 1 NH4, assuming no net increase in nitrifier biomass) for the M1 and D1 cascades would have averaged 9.1 and 9.8 g m − 2 d − 1, respectively, compared to 2.6 and 2.9

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g m − 2 d − 1 for the D2 and D2A cascades (Table 2). Estimates of oxygen release rates for emergent wetland species (many of which are based on questionable assumptions) range from 0 to 12 gO2 m − 2 d − 1, with most in the range of 0.5− 6 gO2 m − 2 d − 1 (Table 10-2, Kadlec and Knight, 1996). Direct diffusion of oxygen through the water surface of the cascade wetlands can be estimated (Kadlec and Knight, 1996, p. 300) to supply a further 0.11 gO2 m − 2 d − 1 (assuming flow velocity=areal hydraulic loading divided by media porosity), which is similar to what would be predicted from molecular diffusion alone. The burrowing and pumping activities of benthic macroinvertebrates, which tend to be more abundant in planted than unplanted SSF wetland media (Tanner; unpublished data), may enhance the rate of surficial oxygen transfer and its penetration into the bed (McCall and Tevesz, 1982). It is likely that nitrifiers would be subject to considerable competition for available oxygen fluxes from heterotrophs (with significantly higher oxygen affinity and energy yield, and thus faster growth rates), as well as CH4, CO, H2S (and other reduced compound) oxidisers (Adams et al., 1996). This makes it doubtful that oxygen fluxes would be sufficient to support the inferred rates of ‘classical’ autotrophic nitrification required to account for the observed mass losses of organic and NH4-N from these systems. Another approach when considering the overall oxygen requirement of coupled nitrification –denitrification is to take into account the oxidising power or ‘oxygen equivalents’ recovered when NOx -N is reduced in the denitrification process (USEPA, 1993). Applying stoichiometric equivalence, the reduction of 1 g NO3-N in denitrification is equivalent to the reduction of 2.86 g of oxygen (USEPA, 1993). This ‘recycling’ of oxygen equivalents can act to decrease the overall oxygen demand for NH4-N conversion to dinitrogen, from 4.57 to 1.67 gO2 g − 1 N − 1(63% decrease). This would lower the average additional net oxygen requirement for N removal in the cascades to  4 g m − 2 d − 1 for M1 and D1, and  1 g m − 2 d − 1 for D2 and D2A (Table 2). Such rates could more realistically be supplied though plantmediated oxygen release, although rates of over 7

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g m − 2 d − 1 required in some stages of the cascades would still be in the upper range measured for wetland plants (Kadlec and Knight, 1996).

3.10. Potential alternati6e mechanisms of coupled nitrification –denitrification Most of what we know about microbial N transformations derives from specialised laboratory studies or mechanised biological nutrient removal plants where simple unitary processes, occurring sequentially in space or time, are optimised by retention of adapted biomass, supply of required electron accepters (e.g. oxygen through mechanical aeration) and donors (e.g. carbon substrates), and, in the laboratory, use of specialised media. Most studies of NH4-N removal in constructed and natural wetlands assume the occurrence of the same ‘classical’ sequence of autotrophic nitrification (NH3 ( “NH2OH) “ NO2 “NO3) followed by respiratory denitrification (NO3 “ NO2 “NO “N2O “N2). However, nitrification –denitrification and organic carbon removal are closely coupled in treatment wetlands and N transformation intermediates (such as NO2 and NO3) rarely accumulate. There is increasing evidence that in such oxygen-limited environments nitrification, denitrification and other microbial processes (e.g. methane oxidation) may be much more closely coupled (also described as integrated or simultaneous) and may include a range of alternative and co-metabolic pathways (Bodelier et al., 2000; Kuai and Verstraete, 1998; Robertson and Kuenen, 1992; Van Loosdrecht and Jetten, 1998). These pathways offer the potential for short-circuiting the classical nitrification – denitrification process. Examples of potential alternative pathways with reduced overall oxygen requirements that have relevance to treatment wetlands include: “ Oxygen-limited autotrophic nitrification – denitrification. Ammonia oxidation via hydroxylamine to nitrite (without further oxidation to nitrate), linked directly to denitrification via nitrous oxide and/or nitrous dioxide (NH4 “ NH2OH “NO2 ( “ NO) “N2O “N2). Omitting the oxygen equivalents of denitrification,

this process would consume 1.7 gO2 g (NH4N) − 1 denitrified. “ Anaerobic ammonium oxidation (ANAMOX). Ammonia oxidation (or nitrite reduction) to hydroxylamine (NH4 (or NO2)“ NH2OH), which is coupled with ammonium to produce hydrazine, then further oxidised to dinitrogen gas (NH2OH + NH4 “ N2H4 “ N2). Omitting the oxygen equivalents of denitrification, this process would consume 1.9 gO2 g (NH4-N) − 1 denitrified. Fdz-Polanco et al. (2001) have recently also reported simultaneous ANAMOX and sulphate reduction resulting in dinitrogen production. “ Heterotrophic nitrification. Oxidation of ammonium by heterotrophs deriving energy from organic substrates. This can be linked directly with denitrification within the same organism; e.g. strains of Paracoccus (formerly Thiosphaera) pantotropha and Pseudomonas stutzeri. Further studies are required to explore the practical importance of such processes in SSF treatment wetlands.

4. Conclusions Cascade mesocosms provide a practical means of studying contaminant removal gradients along treatment wetlands. Cascades receiving wastewaters with differing characteristics, including ratios of COD:TN and balances of nitrogen species, showed contrasting nitrogen process gradients. Emissions of CH4 were recorded for all cascades, particularly the two upstream stages of M1, demonstrating the occurrence of anaerobic catabolism right along the treatment gradient. However, concurrent CO2 emissions were generally around an order of magnitude higher. Overall net plant N uptake, which is likely to have been elevated in our small-scale, harvested systems compared to full-scale treatment wetlands, represented less than 24% of TN removal in the M1, D1 and D2A and 40% in D2 cascades. Contrary to commonly accepted paradigms, nitrification and denitrification occurred concurrently with COD removal, even in upstream

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stages receiving relatively high-strength organicrich wastewaters. Neither net areal nitrification rates, nor first order nitrification rate constants (kA) showed clear correlations with COD concentrations or removal rates. However, denitrification rates closely mirrored net nitrification rates, which showed a unimodal association with average ammonium concentrations in the cascade tanks. Although kA for N mineralisation, nitrification and total N removal were highest for the partially pre-nitrified wastewater tested, considerably higher areal mass removals were recorded in cascades receiving higher strength M1 and D1 wastewaters. The oxygen demand required to support full nitrification of ammonia and mineralised Org-N was in the upper range of that expected to be able to be supplied by plant root-zone oxygen release. Oxygen demand requirements became more realistic if the ‘oxygen equivalents’ potentially recovered in coupled nitrification –denitrification were taken into account. However, alternative coupled or co-metabolic pathways with lower oxygen requirements may also be important for such systems receiving organic and NH4-N-rich organic wastewaters. Acknowledgements This study was funded under Contract CO1X0010 of New Zealand Foundation for Research Science and Technology. Collaboration was facilitated by a grant from NIWA’s Visiting Scientist Programme. NIWA’s Analytical Chemistry Laboratories in Hamilton and Christchurch performed the water, sediment, gas and plant analyses. We are grateful to Albert Van Oostrom, previously of MIRINZ, for valuable conceptual input and Philip Rogers for technical input during the developmental stages of the project. References Adams, D.D., Seitzinger, S.P., Crill, P.M. (Eds.), 1996. Cycling of reduced gases in the hydrosphere. International Association of Theoretical and Applied Limnology, Stuttgart, Germany, p. 203 Communication No. 25.

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