Methane oxidation in vertical flow constructed wetlands and its effect on denitrification and COD removal

Ecological Engineering 128 (2019) 77–88

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Methane oxidation in vertical flow constructed wetlands and its effect on denitrification and COD removal

T

⁎

Thomas Schalka, , Johannes Effenbergerb, Alexander Jehmlichc, Jens Nowakd, Heribert Rustiged, Peter Krebsa, Volker Kühnb a

Institute of Urban and Industrial Water Management, Technische Universität Dresden, 01062 Dresden, Germany Stadtentwässerung Dresden GmbH, Scharfenberger Straße 152, 01139 Dresden, Germany c Erzgebirge Trinkwasser GmbH, Rathenaustraße 29, 09456 Annaberg-Buchholz, Germany d AKUT Umweltschutz Ingenieure Burkard und Partner, Wattstraße 10, 13355 Berlin, Germany b

A R T I C LE I N FO

A B S T R A C T

Keywords: VF wetlands Septic tanks Methane emission Methane coupled denitrification Per-capita methane formation Intermediate

Most types of subsurface flow constructed wetlands include pretreatment units for separation of particulate matter. Due to long retention times of the settled sludge in pretreatment units of small constructed wetlands (CWs), anaerobic degradation processes occur and cause methane emissions into the environment. To minimize methane emissions, small-scale trials were carried out, in order to investigate whether vertical subsurface flow CWs (VF wetlands) are suitable for removal of methane and whether it can be utilized to improve the denitrification rate of VF wetlands. The results show that methane was completely removed for surface loads of up to 19 g CH4/(m2·d). Based on mass balancing, 2% of the fed methane was oxidized with nitrate while 22% were oxidized with oxygen. The major part was attributed to biomass build-up. A correlation between effluent COD load and removed methane load was observed. Published literature indicates that COD accumulation is a result of production of extracellular polymeric substances. Low methane loads have only a slight impact on COD accumulation. It is expected that typical methane loads resulting from pre-sedimentation units may cause effluent COD concentrations to increase by up to 20% for municipal wastewater treatment with VF wetlands. Hence, combined wastewater treatment and methane oxidation was successfully demonstrated, but enhanced denitrification could not be achieved.

1. Introduction Constructed wetlands (CWs) are classified into free water wetlands and subsurface flow wetlands including horizontal and vertical flow systems (Kadlec and Wallace, 2009). The different conditions in these systems lead to different efficiencies particularly for nitrification and denitrification (Vymazal, 2007). Vertical subsurface flow constructed wetlands (VF wetlands) are appropriate for COD removal and thorough nitrification (von Felde and Kunst, 1997; Cooper, 1999). Due to favorable oxygen conditions, denitrification processes within the filter media are of minor relevance and generally amount to 20–40% of total nitrogen (Ntotal) (von Felde and Kunst, 1997) or 20–30% of removed total Kjeldahl nitrogen (TKN) (Fehr et al., 2003). For improved denitrification, downstream and upstream processes are feasible. With upstream denitrification, approximately 50–70% of Ntotal can be removed (Laber et al., 1997). While upstream processes are easy to realize by recirculating effluent into the

⁎

pretreatment unit (Laber et al., 1997; Platzer, 1999), downstream or two-step processes are more complex. These require larger surface areas, because they combine VF wetlands with downstream horizontal flow subsurface wetlands (HF wetlands) (Cooper et al., 1999; Platzer, 1999; Rustige and Platzer, 2001; Vymazal, 2007). Improved nitrogen elimination can be achieved with two-stage VF wetlands, if the first stage is operated with an impounded drainage layer (Langergraber et al., 2011). The first stage of these plants consists of fine gravel and the second stage of coarse sand (DWA, 2017). Compared to conventional VF wetlands, the total surface area required can thus be halved to 2 m2/pe (Langergraber et al., 2011; DWA, 2017), while at the same time significantly increasing the denitrification capacity to > 60% (Langergraber et al., 2011). The denitrification capacity is comparable to upstream processes, but recirculation into the pretreatment unit is not necessary. Single-step denitrification processes may be suitable to overcome the need for recirculation, two-step-processes or the large surface

Corresponding author. E-mail address: [email protected] (T. Schalk).

https://doi.org/10.1016/j.ecoleng.2018.12.029 Received 25 August 2018; Received in revised form 16 December 2018; Accepted 26 December 2018 0925-8574/ © 2018 Elsevier B.V. All rights reserved.

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Fig. 1. Experimental set-up.

et al., 1995; Thalasso et al., 1997; Houbron et al., 1999; Modin et al., 2007). The nature of the intermediate is not clear. According to various authors, possible substances may include methanol (Amaral et al., 1995; Modin et al., 2007; Osaka et al., 2008), acetate (Costa et al., 2000; Modin et al., 2007), citrate (Rhee and Fuhs, 1978) or formaldehyde (Osaka et al., 2008). However, Thalasso et al. (1997) excluded citrate, methanol and formaldehyde as intermediate. Accordingly, Werner and Kayser (1991), Houbron et al. (1999) and Costa et al. (2000) assess denitrification via formaldehyde as less probable. A possible reason for the differing outcomes may be that several types of methanotrophs exist (type I, type II, type X), which assimilate formaldehyde in different ways (Amaral et al., 1995; Hanson and Hanson, 1996). It is possible that experimental conditions (e.g. available oxygen) influence the intermediate (Costa et al., 2000). In addition, a fast assimilation of the intermediate decreases its concentration to such a low level that it is not detectable. Nevertheless, several studies show that denitrification with methane can be realized in one-stage processes under oxygen-limited conditions (Werner and Kayser, 1991; Lee et al., 2001; Islas-Lima et al., 2004; Khin and Annachhatre, 2004). The studied techniques include columns for improving methane solubility (Sollo et al., 1976; Mason, 1977; Rajapakse and Scutt, 1999; dos Santos et al., 2004; Khin and Annachhatre, 2004), down-flow hanging sponge reactors (Hatamoto et al., 2010), trickling filters (Victoria and Foresti, 2011), aerobic biofilm reactors (Modin et al., 2007; Modin et al., 2008) with separated methane and oxygen supply (Waki et al., 2005; Waki et al., 2008) and biofilters (Sly et al., 1993; Melse and van der Werf, 2005; Gebert and Gröngröft, 2006; Park et al., 2009). In general, the application of these processes for treating methane generated in pretreatment units of small CW is not economically feasible, due to the effort for implementation and operation, thus, reducing the benefits of CW (low-energy and low-maintenance operation). In this context, only a simple and robust technology is sensible. The occurrence of methane oxidizing bacteria in CW creates the possibility to use suitable CW for combined wastewater treatment and methane oxidation. VF wetlands are appropriate for methane-coupled denitrification

requirements for combinations of VF and HF wetlands. A possible solution is the use of methane as electron donor for denitrification. Methane is emitted by pretreatment units of small sewage works, due to anaerobic degradation of settled primary sludge as well as anaerobic processes in CW. The biogas formation potential and the degradation rate of the settled volatile suspended solids (VSS) are affected primarily by the wastewater temperature and the sludge retention time (SRT) (see Metcalf and Eddy, 1991). The emission from CW depends on the temperature, the loading and the construction type (Mander et al., 2014). Methane is formed in free water wetlands (Inamori et al., 2003; Johansson et al., 2004) and in subsurface flow wetlands, particularly in the inflow section of HF wetlands (Mander et al., 2005; Søvik et al., 2006; Zhu et al., 2007; Picek et al., 2007) and in HF wetlands operating with continuously high moisture content (Nurk et al., 2005). Methane formation potential attributed to VF wetlands is demonstrably smaller (Mander et al., 2014). Considering the published data, favorable conditions for preventing methane emissions are provided by planted VF wetlands with intermittent loading and resting periods (Inamori et al., 2007; Wang et al., 2008). Besides methanogenic microorganisms, bacteria with the ability to utilize nitrate for methane oxidation occur in CW (Heilmann and Carlton, 2001; Inamori et al., 2003; Zhu et al., 2007; Wang et al., 2008). The simplified reaction equation:

5 CH 4 + 8 NO−3 → 4 N2 + 8 OH− + 6 H2 O+ 5 CO2

(1)

is used by several authors including Harremoës and Henze Christensen (1971), Thalasso et al. (1997), Islas-Lima et al. (2004), and Modin et al. (2007). The actual reactions are more complex because different microorganism groups are involved in this process (Thalasso et al., 1997). Aerobic microorganisms oxidize methane via the intermediates methanol, formaldehyde and formate (Whittenbury et al., 1970; Dalton and Leak, 1985). It is likely that oxidation of ammonia with coupled denitrification of nitrate is based on interactions between methanotrophic and denitrifying bacteria (Zellner et al., 1995; Thalasso et al., 1997; Costa et al., 2000). Methanotrophs assimilate methane and produce an organic intermediate utilized by denitrifying bacteria (Amaral 78

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2. Methods

162 d (stage 2B). The recirculation ratio (Qrecirculation/Qwastewater) was set to 0.5. The wastewater was pumped twice per day in stages 1 and 2A and three times per day in stage 2B from the storage tanks for raw wastewater and the recirculation tanks. Recirculation and feeding pumps operated in parallel in stage 2B. The wastewater feed and the methane feed were synchronized. Wastewater inflow started when the first half of the current methane aliquot was fed.

2.1. Examination of methane build-up in septic tanks

2.3. Wastewater used and seeding sludge

To investigate the methane formation potential of septic tanks, three small-scale closed, stainless-steal reactors (ST10, ST20, ST30), each with a volume of 17 L (Fig. 1B), were operated for 531 days. Reactor temperatures were maintained continuously at 10 °C, 20 °C, and 30 °C with two combined heating and cooling thermostats (LAUDA Alpha RA 8 – Lauda Dr. R. Wobser GmbH & Co. KG, Lauda-Königshofen, Germany) and one heating thermostat (EFM1 – Funke Medingen, Tharandt, Germany). The reactors were seeded with mesophilic digested sludge (each 1 L with each 30 g TSS, 17 g VSS, 29 g COD), undigested primary sludge (each 3 L with each 78 g TSS, 57 g VSS, 76 g COD), and raw wastewater (each 12 L with each 5.6 g TSS, 10 g COD) of the Dresden wastewater treatment plant (see Section 2.3). The hydraulic retention time (HRT) of the supernatant amounted to at least 10 d. The operating conditions were approximated to the design rules for pretreatment units of small sewage works according to the German standard DIN 4261-1 (DIN, 2010). Each reactor was fed on weekdays with 2.3 L of raw wastewater (see Section 2.3). To avoid aeration, the raw wastewater was fed beneath the water surface with an immersed tube. Before feeding, supernatant was withdrawn with a siphon.

Municipal wastewater of the wastewater treatment plant (WWTP) of Dresden (design capacity: 740.000 PE, combined sewer system, coordinates: 51.072816/13.677884) was used. Feed samples were obtained from the WWTP after coarse solids removal in a multi-level screening system (bar spacing: 1st step: 65 mm, 2nd step: 15 mm) and sand removal in an unaerated grit chamber. Primary sludge and digested sludge for seeding the small-scale septic tanks were taken from Dresden WWTP too. Undigested primary sludge was obtained after gravity thickening, digested sludge from a sludge line of the digester effluent (2 × 10,500 m3, HRT = 21 d, T = 35 °C, coordinates: 51.077782 / 13.681729).

due to their high potential for nitrification and their low potential for methane formation. Hence, a VF wetland was adapted for a single-step nitrification and methane-coupled denitrification process. Additionally, methane production of pretreatment units of small sewage works was investigated.

2.4. Sampling and analytical methods For examination of methane oxidation in the small-scale columns, daily samples (inflow: nstage_1 = 222, nstage_2A = 60, nstage_2B = 115, effluent (each column): nstage_1 = 222, nstage_2A = 60, nstage_2B = 115) were taken from the effluent and the inflow tanks and added to weekly composite samples (inflow: nstage_1 = 47, nstage_2A = 12, nstage_2B = 24, effluent (each column): nstage_1 = 47, nstage_2A = 12, nstage_2B = 24). To determine each flow, the used tanks were weighed. All samples were analyzed according to the German standard methods for examination of water, wastewater and sludge. The inflow samples were analyzed for total chemical oxygen demand (CODtotal, tube tests according to the German standard DIN ISO 15705 – Hach Lange GmbH, Germany), dissolved COD (CODdis, tube tests according to the German standard DIN ISO 15705 – Hach Lange GmbH, Germany) after filtration through 0.45 µm membrane filters (Sartorius AG), total suspended solids (TSS, German standard DIN 38409-2), total Kjeldahl nitrogen (TKN, German standard DIN EN 25663 – BÜCHI Labortechnik GmbH, Germany) and ammonium-nitrogen (NH4-N, spectroquant 114752 – Merck KGaA, Germany). Effluent samples were additionally analyzed for nitrate (NO3-N, German standard DIN EN 38405-9) nitrite (NO2-N, spectroquant 114776 – Merck KGaA, Germany) and in the last thirteen weeks of the trials for dissolved organic carbon (DOC) according to the German standard DIN EN 1484 by using a TOC-V CPH analyzer (Shimadzu Deutschland GmbH). Furthermore, temperature, and pH were measured in all storage tanks (pH 323 with SenTix 21, WTW Wissenschaftlich-Technische Werkstätten GmbH, Germany). To investigate possible intermediates generated during methane oxidation, effluent samples were analyzed for formic acid, acetic acid, citric acid, hydroxybutyric acid by ion chromatography (Dionex ICS-3000, Dionex GmbH, Germany) and for formaldehyde by tube tests (Hach Lange GmbH, Germany). For determination of methane degradation, both columns were covered once a week for 24 h. After this period, the gas quality was measured in the headspace between the top of the filter bodies and the column covers. Additional gas quality measurements were conducted randomly in the gas-feed-pipes. For gas analyses a portable gas detector was used (LMSxi, GasData Ltd., Whitley, Coventry (UK)). Gas analyses and methane balancing began four months after initial methane dosing. Regarding the small-scale septic tanks, daily samples (n = 359) were taken on weekdays from the supernatant and from the inflow. Daily samples were unified to weekly composite samples (inflow: n = 76, supernatant (each reactor): n = 76). All samples were analyzed for total COD and TSS. Furthermore, temperature, and pH were

2.2. Small-scale methane oxidation tests For studying methane oxidation in VF wetlands, two small-scale columns were used (Fig. 1A). The reference column (REF) was operated without methane dosing for the entire experimental duration of 610 days. The second column (MET) was similarly operated for 358 days and subsequently fed with methane for 252 days (100% CH4, Air Liquide Deutschland GmbH, Düsseldorf (Germany)). For methane dosing, a timer-controlled TELAB BF 414/1002S laboratory diaphragm dosing pump was used (Telab UG, Solingen (Germany)). The dosing rate of the gas-feeding pump was 30 mL/min. The columns were made of polyethylene (type: SB/LB-300, ARICON Kunststoffwerk GmbH, Solingen (Germany)) with a diameter of 660 mm and a surface area of 0.34 m2. They were filled with sand (grain size: 0–2 mm, h = 600 mm) as filter media and gravel as drainage layer media (grain size: 4–8 mm, h = 200 mm). A perforated pipe for gas feeding was installed (d = 40 mm) at half the height of the sand filter layers of each reactor. Two ball valves permitted sealing of the pipes. The gas feed was connected to the first ball valve. The second ball valve at the end of the pipe was used for taking gas samples. The column feed was prepared from mechanically treated municipal wastewater (see Section 2.3). The feed was further treated by sedimentation in a batch-wise operated small-scale sedimentation tank (V = 300 L, HRT = 2 h). Timer-controlled peristaltic pumps (two parallel pumps per column) were used to feed the reactors. The treated wastewater was collected in an effluent tank. Every column was equipped with separate tanks for feed, effluent and recirculation (each 35 L). These tanks were open to gas exchange with the surrounding air. All pumps used were identical (PD 5006 SP standard; Heidolph Instruments GmbH & Co.KG Schwabach (Germany)). Column operation was divided in three stages. During the first 358 d, the columns were operated without recirculation of biologically treated wastewater and without methane feed (stage 1). After that, column MET was operated without recirculation but with methane feed for 90 d (stage 2A). Lastly, both columns were operated with effluent recirculation and column MET was additionally fed with methane for 79

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Table 1 Operating conditions and removal of COD and TSS of lab-scale septic tanks dependent on the temperature. Reactor No.

T

Q

Inflow

Total inflow

Daily inflow

Loading

Effluent

Removal rate

Concentration

Concentration

Discharged sludge after 271 days

after 566 days

[−]

[°C]

[L]

[L/d]

CODtotal [g/(m3∙d)]

TSS [g/(m3∙d)]

CODtotal [mg/L]

TSS [mg/L]

CODtotal [mg/L]

TSS [mg/L]

CODtotal [%]

TSS [%]

TSS [g/L]

VSS [% TSS]

TSS [g/L]

VSS [% TSS]

1 2 3

10 20 30

835 835 835

2.3 ± 0.3 2.3 ± 0.3 2.3 ± 0.3

46 ± 14 46 ± 14 46 ± 14

31 ± 11 31 ± 11 31 ± 11

483 ± 133 483 ± 133 483 ± 133

326 ± 105 326 ± 105 326 ± 105

164 ± 61 113 ± 28 97 ± 20

97 ± 43 72 ± 20 60 ± 17

66 77 80

70 78 82

18.8 11.7 10.9

56 50 46

17.4 16.6 19.2

54 48 45

the calculated load of the dissolved CH4, g; CODww,eff. is the COD effluent load, g; and CODww,rem. is the COD removed from the wastewater, g. The gas substances were calculated according to Henry's Law and the equations given by Svardal (1991):

measured. Additionally, TSS and VSS of the sludge were analyzed according to the German standard DIN EN 12879. 2.5. Balancing For mass balancing, the term degradation addresses the completely decomposed part of the fed substances. The term removal addresses the sum of the completely decomposed fraction and the fraction that is converted into biomass and intermediates. For assessing the impact of methane removal on wastewater treatment, COD removal is calculated based on the COD of the wastewater, without impacts of the fed methane. This is a simplified approach that does not address whether the effluent load results from incomplete degradation of the feed (i.e. overload) or production of biomass or biomass products built up during wastewater treatment and methane degradation. This approach was chosen to illustrate the impact of methane feeding on wastewater effluent concentrations. Nitrification rate is related to the inflow TKN. TKN removal includes the nitrified nitrogen and the nitrogen demand for biomass growth. Nitrogen uptake by biomass was estimated as 2.5% of the fed wastewater COD including the COD equivalent of the fed methane. This nitrogen uptake was chosen in accordance to the German standard ATV-DVWK-A 131 (2000). However, the nitrification rate will be overestimated by this approach, if biomass nitrogen content and/or biomass yield relative to COD input are actually higher than the respective assumptions of the German standard ATV-DVWK-A 131 (2000). Biomass yield could not be measured directly but had to be estimated as the result of COD balancing. There are two options for balancing the COD. Both are black box approaches, in which biomass growth associated with methane oxidizing microorganisms (MOM) is calculated as difference between inflow and outflow data. In the first approach, biomass growth results from the difference between CH4 feed and CH4 losses by oxidation, breakthrough and solution processes (Eq. (2)):

cCH4 = MCH4 ·p·K H,CH4·yCH4 lgK H,CH4 =

(5)

where CCH4 is the saturation concentration, g/L; MCH4 the molar mass, g/mol; p the pressure, bar; K H,CH4 the Henry constant, mol/(L⋅bar); yCH4 the mole fraction, T the temperature, K; A, B and C are constants (A CH4 = 2, 370.4 ; BCH4 = 16.33; CCH4 = 0.0185). 3. Results and discussion 3.1. Methane generation in septic tanks The loading rate of the reactors averaged 46 ± 14 g CODtotal/ (m3∙d). Inflow concentrations (Table 1) fluctuated depending on the inflow conditions of the Dresden WWTP. Compared to full-scale plants (e.g. see Ulrich et al., 2009), the efficiency of the reactors was higher and increased with rising temperatures (Table 1). Batch-operation may have led to undisturbed sedimentation, which benefitted the efficiency. Sludge was removed when sludge particles were observed in the withdrawn supernatant. This occurred on day 271 and on day 566 (end of the experiment). In both cases, the VSS content of the digested sludge was comparable. The values decreased slightly with increasing SRT. The VSS degradation increased with rising temperature from 42% at 10 °C to 59% at 30 °C after a SRT of 566 days (Table 2). Due to the long SRT, influences of the seeding sludge are negligible. Estimation of methane formation needs to account for collectible methane gas as well as the CH4-loss caused by leakages and solution processes. These losses are significant due to the low absolute methane generation potential resulting from the low organic loading rate and the use of municipal wastewater as feed medium. Hence, methane production is estimated based on degradation of COD and VSS. Assuming a per-capita (p) primary sludge amount of 28 g TSS/(p∙d), a VSS/TSS ratio of the raw primary sludge of 0.75 (DWA, 2014) and a typical SRT of 0.75–1.0 years, mean theoretical methane production for temperatures between 10 and 30 °C can be estimated to 5.1–7.3 L CH4/(p∙d) using the stoichiometric methane production of 250 mg CH4/g CODdegraded. Depending on the temperature, a part of the formed methane is solved in water and washed out with the effluent. Methane emissions from small sewage works vary strongly (for some examples see Table 3). On the one hand, temperature is of major significance for anaerobic degradation processes in the settled sludge. On the other hand, methane emissions are affected by the SRT and the desludging interval, by methane solution processes in the wastewater (HRT of the wastewater) as well as the gas exchange at the boundary surface between air and water. Onsite-measurements only allow a short overview of the possible concentration range.

CODMOM,app1 = CODCH 4,feed − CODCH4 ,ox − CODCH4 ,HS − CODCH4 ,dis. (2) where CODMOM,app1 is the COD of the methane oxidizing bacteria, g; CODCH4,feed is the COD equivalent load of the fed CH4, g; CODCH4,ox. is the COD equivalent load of the oxidized CH4 (see Section 3.2.4), g; CODCH4,HS is the COD equivalent load of the CH4 contained in the headspace, g; and CODCH4,dis. is the calculated load of the dissolved CH4, g. The second approach involves the COD contained in the wastewater by assuming the COD removal efficiency of the reference column to the MET column (Eq. (3)):

CODMOM,app2 = CODCH4 ,feed + CODww,feed − CODCH4 ,ox − CODCH4 ,HS − CODCH4 ,dis − CODww,eff. − CODww,rem

A CH4 − BCH4 + CCH4 ·T T

(4)

(3)

where CODMOM,app2 is the COD of the methane oxidizing bacteria, g; CODCH4,feed is the COD equivalent load of the fed CH4, g; CODww,feed is the COD load of the fed wastewater, g; CODCH4,ox. is the COD equivalent load of the oxidized CH4 (see Section 3.2.4), g; CODCH4,HS is the COD equivalent load of the CH4 contained in the headspace, g; CODCH4,dis. is 80

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Table 2 Degradation of total COD and VSS and the methane forming capability of settled sludge dependent on the temperature. Reactor No.

T

COD balance

COD degradation

VSS balance

CH4 formation4

VSS degradation

Inflow1

Effluent2

Reactor3

Absolute

Relative

Inflow1

Effluent2

Reactor3

Absolute

Relative

[−]

[°C]

[g]

[g]

[g]

[g]

[%]

[g]

[g]

[g]

[g]

[%]

[L/(p∙d)]

1 2 3

10 20 30

518 514 519

260 189 168

88 100 75

172 226 276

33 44 53

281 280 282

113 79 68

52 56 18

117 145 166

42 52 59

5.1 6.4 7.3

1 2 3 4

Seeding sludge + sewage. Supernatant + desludging after 271 days (5 L each reactor). Sludge + supernatant. Calculated.

0.9–2.6 g CH4/(m2∙d). Tested conditions covered a wider range, in order to determine the critical load for a methane breakthrough.

3.2. Methane oxidations trials 3.2.1. Wastewater load The hydraulic wastewater surface load was approximately constant in all stages and averaged 34–35 L/(m2∙d). Differences were caused by defects of the tubes used for wastewater pumping. The applied hydraulic surface load did comply with the design per-capita wastewater amount for small sewage works in Germany (150 L/(p∙d)). In stage 2B, an additional hydraulic load of approximately 6 L/(m2∙d) resulted from the effluent recirculation (Table 4). Since the hydraulic wastewater load was constant within the stages, the COD and TKN loads did only depend on wastewater concentrations. The variation of COD concentrations and corresponding COD loads (Table 4) were greater than respective variations for TKN (Table 5). Average loads were 8.5–9.7 g/(m2∙d) for COD and 1.8–1.9 g/(m2∙d) for TKN. For comparison, the surface loads resulting from per-capita loads after sedimentation (72 g CODtotal/(p∙d) and 9.9 g TKN/(p∙d) for HRTs > 2.5 h) (ATV-DVWK, 2003; DWA, 2017) and the design surface area required for small VF wetlands (4 m2/p for one-step sand filters) (DWA, 2017) can be calculated to 18 g CODtotal/(m2∙d) and 2.5 g TKN/ (m2∙d), respectively. Due to the chosen hydraulic load, actual CODtotal and TKN surface loads were lower in the trials than the design values.

3.2.3. Removal efficiency without methane feeding (stage 1 and 2A) Both columns were operated in parallel for one year before methane feed to column MET started. The removal rate for CODtotal was 95%. The major part of the effluent COD was dissolved (see Table 4). The effluent concentrations regarding CODtotal averaged 14 ± 3.9 mg/L (REF) and 14 ± 4.2 mg/L (MET). The TKN removal averaged 2.0 g TKN/(m2∙d), while the nitrification rate was 1.7 g /(m2∙d) or 87% (see Table 5). In general, NH4-N concentrations in both effluents did not exceed 0.5 mg/L and NO3-N concentrations were 46 mg/L on average. The average denitrification rate of 0.1 ± 0.3 g/(m2∙d) or 5% was small. However, some equalization of concentrations and respective process rates was caused by the use of weekly composite samples. After start up, a temporary nitrite accumulation occurred, which lasted for four weeks. Peak values were 7.6 and 7.7 mg NO2-N/L and occurred after three weeks (data not shown). As the samples were taken as weekly composite samples, it is obvious that the absolute peak values were higher than the measured values. Temporary nitrite accumulation can occur during running-in of nitrification due to adaption of ammonia oxidizing microorganisms and nitrite oxidizing microorganisms to high ammonia loads (Nowak et al., 1999). The adaption period depends on the operational conditions and can last several weeks (Nowak et al., 1999). Without methane feeding (stage 1), the pH of both columns dropped from 8.2 after start-up to 4.0–4.5 within a year. Since no methane was fed to column REF in stage 2A and stage 2B, this level remained for the nine following months until the trials ended. No adverse effects on nitrification rates were observed. On the one hand, the ammonia effluent concentrations were stable at a very low level (Table 5), on the other

3.2.2. Methane load The methane feed to column MET was gradually increased from 0.54 L CH4/d to the peak load of 12.6 L CH4/d. This corresponded to a surface load of 1.0–27 g CH4/(m2∙d) (Fig. 2). Based on the calculated methane generation of pretreatment units of small sewage works (5.1–7.3 L/(p∙d), see Section 3.1) and surface requirements of 2 and 4 m2/p, for different types of VF wetlands, the potential methane surface loading for VF wetlands with pre-sedimentation is approximately Table 3 CH4 and CO2 concentrations in the headspace of pretreatment units. WWTP

Treatment process1

Capacity [p]

Septic tank2

Wastewater origin

Buffer tank2

1st chamber

A

Septic tank – buffer tank – VF wetland

80

Camping site

B

Septic tank – buffer tank – SBR

160

Camping site

C

Septic tank – buffer tank – VF wetland

10

Holiday lodge

D

Septic tank – VF wetland – HF wetland

145

Village

Mean n Mean n Mean n Mean n

1

2nd chamber

CH4 [Vol-%]

CO2 [Vol-%]

CH4 [Vol-%]

CO2 [Vol-%]

CH4 [Vol-%]

CO2 [Vol-%]

1.8 3 0.1 4 0.7 10 0.3 3

0.2 2 0.9 4 0.1 8 0.3 3

0.5 2 0.0 3 0.0 8 0.1 2

0.1 1 0.4 ± 0.4 3 0.1 ± 0.1 8 0.1 ± 0.1 2

0.1 ± 0.2 3 0.0 ± 0.0 5 0.0 ± 0.0 8 – –

0.2 ± 0.1 2 0.8 ± 0.3 5 0.1 ± 0.2 7 – –

± 2.8 ± 0.2 ± 1.5 ± 0.4

± 0.2 ± 1.4 ± 0.2 ± 0.2

± 0.1 ± 0.0 ± 0.0 ± 0.1

A description of the WWTPs is contained in Schalk (2017). Onsite-measurements with a portable gas detector (LMSxi, GasData Ltd., Whitley, Coventry (UK)) and a gas detection device (DRÄGER X-am 2000, Drägerwerk AG & Co. KGaA Lübeck (Germany)). 2

81

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Table 4 Operational parameters of the columns tested. Stage1

1 2A 2B

Column

REF MET REF MET REF MET

Inflow

Effluent

Wastewater [L/(m2∙d)]

pH [−]

T [°C]

CODtotal [g/(m2∙d)]

CODtotal [mg/L]

CODdissolved [mg/L]

pH [−]

35 34 34 34 35 35

7.8 ± 0.3

18 ± 2.4

283 ± 98

169 ± 64

7.9 ± 0.2

15 ± 1.8

248 ± 63

150 ± 42

7.9 ± 0.2

19 ± 2.9

9.9 9.7 8.5 8.5 9.4 9.6

259 ± 53

148 ± 39

5.8 5.6 4.2 4.2 4.3 5.9

± ± ± ± ± ±

5.6 5.9 1.1 1.7 1.7 1.7

± ± ± ± ± ±

3.5 3.5 2.1 2.2 1.7 1.8

± ± ± ± ± ±

Recirculation T [°C]

1.2 1.2 0.1 0.1 0.2 1.0

20 20 18 17 21 21

CODtotal [mg/L]

± ± ± ± ± ±

1.9 2.0 0.4 0.4 2.8 2.9

13.8 13.9 10.7 11.5 14.0 33.6

± ± ± ± ± ±

3.4 3.9 1.1 1.4 3.1 23

CODdissolved [mg/L]

[L/(m2∙d)]

12.3 ± 3.0 12.3 ± 3.3 9.4 ± 1.3 9.6 ± 1.6 12.0 ± 2.6 25.6 ± 15

– – – – 5.8 ± 0.8 5.9 ± 0.7

1 Stage 1 … without methane feed. Stage 2A … with methane feed but without recirculation (methane load on average: 2.1 ± 0.9 g/(m2∙d), t = 12 weeks). Stage 2B … with methane feed and with recirculation (methane load on average: 13 ± 8.2 g/(m2∙d), t = 24 weeks).

10.0

100 REF - Nitrate [mg/L] MET - Nitrate [mg/L] REF - pH MET - pH MET - Exhaust CH4 [g/(m² d)] MET - CH4 surface load [g/(m² d)]

pH [-] Exhaust CH4 [g/(m² d)]

9.0 8.0 7.0

90 80 70

6.0

60

5.0

50

4.0

40

3.0

Stage 1

30

Stage 2B

Stage 2A

2.0

20

1.0

10

0.0

0 Jan Mar May

3.2.4. Methane removal (stage 2B) Methane was removed completely for surface loads of up to 19 g CH4/(m2∙d). At higher loads, methane broke through the filter and it was detectable in the exhaust air (Fig. 2). The CO2-concentrations in the exhaust air of the reference column averaged 1.3 ± 0.3 vol-% and were attributed primarily to COD removal. The CO2 concentrations for exhaust air of the MET-column were considerably higher at 4.5 ± 1.8 vol-%. The difference between these CO2 concentrations was attributed to the CO2 formation caused by methane oxidation. Fig. 3 presents the correlation between calculated methane removal and the measured CO2 concentrations in the headspace between the filter and the cover of the MET column. Since equivalent CO2 amounts result from oxygen and nitrate metabolism, the increase of CO2 concentrations indicates methane degradation but does not allow conclusions regarding the degradation pathway. The nitrate-coupled methane oxidation was estimated based on the nitrified nitrogen, the recirculated NO3-N load, the effluent NO3-N load and the stoichiometric nitrate demand for methane oxidation (1.6 mol NO3/mol CH4). For balancing, it was assumed that the partial pressure of methane in the pore volume of the filter sand close to the gas inlet pipe was nearly 100% during feeding. This is the worst case with the maximum potential methane loss by methane washout with the

Nitrate [mg/L] CH4 surface load [g/(m² d)]

hand the estimated nitrification rate was constant too. In column REF 87.2 ± 2.8% of the fed TKN was nitrified in stage 1, in stage 2A 88.7 ± 1.5%, and in stage 3 86.8 ± 1.7% (Table 5). This behavior indicates that stable nitrification in acidic conditions is possible at the chosen wastewater loads, if the treatment system prevents negative influences on biomass retention (e.g. floc break-up). It is common that low pH values inhibit nitrification, but there are also studies, which suggest the possibility of nitrification at low pH-values (Yao et al., 2011; Zhang et al., 2012; Stempfhuber et al., 2015; Lehtovirta-Morley et al., 2016). In summary, parallel operation of both columns led to comparable results.

Jul

Sep Nov

Jan Mar May

Jul

Sep

Fig. 2. Operational parameters of both columns.

Exhaust CO 2 [g/(m² d)]

25 20 15 10 5

CO2,exhaust = 0.71 · CH4,removed R² = 0.90

0 0

5

10 15 20 Removed CH4 [g/(m² d)]

25

30

Fig. 3. Carbon dioxide formation during methane removal.

effluent. The losses by carbon dioxide solution and carbonate precipitation processes influence balancing due to the assumption, that only the CO2 difference between both columns represents the oxidation of the fed

Table 5 Nitrogen balancing neglecting start-up operation during the first four weeks. Stage1

1 2A 2B

Column

REF MET REF MET REF MET

Inflow + Recirculation Ntotal

TKN

[g/(m2∙d)]

[g/(m2∙d)]

2.0 2.0 1.9 1.9 2.5 2.4

2.0 2.0 1.9 1.9 1.8 1.8

± ± ± ± ± ±

0.5 0.4 0.3 0.3 0.5 0.6

± ± ± ± ± ±

0.5 0.4 0.3 0.3 0.4 0.4

Effluent

Nitrified

Denitrified

N

N

N

NO3-N

NH4-N

[mg/L]

[g/(m2∙d)]

[g/(m2∙d)]

[mg/L]

[g/(m2∙d)]

[mg/L]

[g/(m2∙d)]

[g/(m2∙d)]

[g/(m2∙d)]

56 56 53 54 50 50

– – – – 0.7 ± 0.2 0.5 ± 0.3

0.0 0.0 0.0 0.0 0.0 0.0

0.3 0.3 0.0 0.1 0.1 0.6

1.7 1.7 1.6 1.5 2.3 1.6

49 49 46 44 45 31

0.2 0.2 0.2 0.2 0.2 0.3

1.8 1.7 1.7 1.7 1.6 1.5

0.1 0.1 0.1 0.2 0.0 0.4

± ± ± ± ± ±

12 11 9.0 8.3 11 11

± ± ± ± ± ±

NO3-N

Biomass2

0.0 0.0 0.0 0.0 0.0 0.1

± ± ± ± ± ±

0.6 0.7 0.0 0.1 0.1 1.1

± ± ± ± ± ±

0.3 0.3 0.3 0.2 0.4 0.8

± ± ± ± ± ±

7.8 8.1 6.0 7.1 6.6 15

± ± ± ± ± ±

0.1 0.1 0.1 0.1 0.0 0.1

± ± ± ± ± ±

0.4 0.4 0.4 0.3 0.3 0.3

± ± ± ± ± ±

0.3 0.3 0.2 0.1 0.3 0.6

1 Stage 1 … without methane feed. Stage 2A … with methane feed but without recirculation (methane load on average: 2.1 ± 0.9 g/(m2∙d), t = 12 weeks). Stage 2B … with methane feed and with recirculation (methane load on average: 13 ± 8.2 g/(m2∙d), t = 24 weeks). 2 Calculated with 0.025 ∙ COD according to the German standard ATV-DVWK-A 131 (2000).

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Table 6 Methane removal paths at high CH4 load conditions. Methane (CH4) Feed

Nitrate (NO3-N) Methane losses

Methane removal

Nitrified/

Total

Effluent

Exhaust

Total

CH4 Oxidation

CH4 [g/(m2·d)]

CH4 [g/(m2·d)]

CH4 [g/(m2·d)]

CH4 [g/(m2·d)]

CH4 [g/(m2·d)]

Via NO3 [g/(m2·d)]

11.4 13.3 15.2 18.9 22.7 26.5 24.6

1.29 1.32 1.20 1.17 1.29 1.75 1.38

1.29 1.32 1.20 1.17 1.17 1.18 1.30

0.00 0.00 0.00 0.00 0.12 0.57 0.08

10.1 12.0 14.0 17.7 21.4 24.8 23.2

0.05 0.09 0.11 0.24 0.42 0.65 0.86

3.2.5. Denitrification With increasing methane load, nitrate concentrations decreased in column MET while pH increased. The differences between the effluent NO3-N concentrations of both columns did correlate with the CH4 surface load: (6)

where CNO3 - N is the nitrate concentration in the reference (REF) and the methane (MET) column, mg/L; and CH4 is the CH4 load (MET), g/ (m2∙d); R2 = 0.93, n = 20. The pH of the reference column remained stable at 4.2 ± 0.2 while the pH of the MET column increased gradually with increasing CH4 load to 7.5 (see Fig. 2). Similarly, the differences between the pH values did correlate with the differences of the nitrate concentrations:

Biomass

Total

Via O2 [g/(m2·d)]

CH4 [g/(m2·d)]

NO3-N [g/(m2·d)]

NO3-N [g/(m2·d)]

NO3-N [g/(m2·d)]

2.71 3.11 3.90 3.87 5.06 5.91 4.34

7.34 8.75 10.0 13.6 15.9 18.2 18.0

2.19 2.04 2.14 2.07 1.81 1.55 1.43

2.12 1.91 1.99 1.74 1.22 0.64 0.23

0.07 0.13 0.15 0.33 0.59 0.91 1.20

1.6

(7)

Denitrified NO3-N [g/(m² d)]

pHMET − pHREF = 0.74·ln(CNO3 - N,REF − CNO3 - N,MET)+0.24

Denitrified

biomass. Nevertheless, the impact of methane oxidation on denitrification is small. The average nitrate removal rate was 0.01–0.07 mol NO3-N/mol CH4. Similar values are given in literature. The denitrification rate published by Werner and Kayser (1991) is 0.02 mol NO3-N/mol CH4; the rate given by Houbron et al. (1999) is 0.06–0.07 mol NO3-N/mol CH4. Modin et al. (2010) achieved 0.07–0.11 mol NO3-N/mol CH4 during headspace trials and 0.25–0.36 mol NO3-N/mol CH4 using a membrane biofilm reactor. Both Werner and Kayser (1991) and Houbron et al. (1999) stated that less than 7% of the fed methane was converted to methanol, which could be utilized for denitrification. In the present study, only 2% of the methane was utilized for denitrification (see Section 3.2.4). For surface loads less than 5 g CH4/(m2∙d) no distinct effects on the operation and the effluent values were observed (Fig. 4). With regards to expected methane surface loads of approximately 0.9–2.6 g CH4/ (m2∙d) resulting from methane formation in pre-sedimentation units (see Section 3.1), the methane-coupled denitrification is irrelevant for reduction of nitrate concentrations. Only CH4 loads > 24 g CH4/(m2∙d) resulted in almost complete denitrification of nitrate (Table 6, Fig. 4). The NH4-N concentrations in column MET began to increase when the CH4 load exceeded 8 g CH4/(m2∙d), but they remained below 1.0 mg NH4-N/L while the CH4 load was less than 23 g CH4/(m2∙d). The concentration values of column REF did not increase. While CH4 loads remained below 8 g CH4/(m2∙d), NH4-N concentrations averaged 0.04 ± 0.02 mg NH4-N/L (n = 21). For higher CH4 loads, the NH4-N concentrations increased on average to 0.29 ± 0.31 mg NH4-N/L (REF: 0.09 ± 0.10 mg NH4-N/L, n = 9). As the CH4 load exceeded 23 g CH4/(m2∙d), the NH4-N concentrations averaged 1.78 ± 1.45 mg NH4-N/L in the MET column (REF: 0.05 ± 0.05 mg NH4-N/L). Even though the NH4-N concentrations did not correlate with the methane load or with the methane in the exhaust air, high methane concentrations appeared to impair the nitrification. It

methane. Hence, the actual oxidized methane load can be underestimated. Up to a surface load of 9.5 g CH4/(m2∙d), the NO3-N load was stoichiometrically sufficient to oxidize the fed methane completely. However, only a small part of nitrate was actually denitrified. Results are given for surface loads > 9.5 g CH4/(m2∙d). At lower load conditions, balancing resulted in several negative denitrification values due to the slight differences between the nitrogen substances. For methane loads > 9.5 g CH4/(m2∙d), mass balancing shows that on average less than 2% of methane was oxidized with nitrate (Table 6). While the removed amount of methane is equivalent to the produced amount of carbon dioxide (Fig. 3), the molar ratio of 3.9 is substantially higher than the stoichiometric ratio of 1.0. It is unlikely that this caused by incomplete measurement of carbon dioxide or solution processes within the filter. Assuming that the measured additional CO2 of the METcolumn is caused by CH4 oxidation, only 24 ± 2.9% of the fed methane was oxidized via oxygen or nitrate. The major part has been converted to an organic substance, e.g. an intermediate or biomass (see Section 3.2.6).

CNO3 - N,REF − CNO3 - N,MET = 1.21·e0.134·CH4

Effluent

where pHMET is the pH value in the MET column; pHREF is the pH value in the reference column; CNO3 - N is the nitrate concentration in the reference (REF) and the methane (MET) column, mg/L; R2 = 0.83, n = 20. The decrease of the nitrate concentrations and the increase of the pH generally indicate denitrification processes. The increased COD loading by methane will also cause increased biomass buildup. This will increase incorporation of nitrogen into biomass and reduce absolute nitrification. Though unlikely, it cannot be ruled out that nitrogen demand of methanotrophic microorganisms for methane oxidation is partly covered by nitrate uptake (see Section 3.2.6). Thalasso et al. (1997) found that up to 100% of formed nitrate was metabolized by

NO3-Ndenitrified = 2 · 10-5 · CH4,removed3.39 R² = 0.95

1.4 1.2 1.0 0.8 0.6 0.4 0.2 0.0 5

10

15 20 Removed CH4 [g/(m² d)]

25

30

Fig. 4. Relation between the removed methane and the denitrification rate. 83

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120

30

detectable in the effluent of the MET column. Respective concentration values were close to corresponding detection levels. The ratio between dissolved COD and DOC of the accumulated substances averaged 2.7 (R2 = 0.93, n = 13), which is close to the ratio of 2.63 given by Thalasso et al. (1997). Based on the results published by Hilger et al. (2000), Chiemchaisri et al. (2001) or Wilshusen et al. (2004), it is likely that COD accumulation is a microbial reaction to the methane exposition, which is characterized by the production of extracellular polymeric substances (EPS). Since methane oxidation is conducted by methanotrophic microorganisms, a part of the fed methane must be utilized for biomass growth. This is indicated by the higher phosphorous demand of the MET column:

COD [mg/L]

100

25

REF - COD effluent concentration [mg/L] MET - COD surface load [g/(m² d)]

80

20

60

15

40

10

20

5

0

0 Jan Mar May Jul Sep Nov Jan Mar May Jul Sep Nov

CH4 surface load [g/(m² d)]

MET - COD effluent concentration [mg/L]

Fig. 5. COD effluent concentrations of both columns.

oPO4 − Pconsumed = 0.008·CH 4,consumed is possible that methane inhibits the NH4-N oxidation directly by suppressing the oxygen required for nitrification but also that the oxygen consumption for methane oxidation competes with the oxygen demand for nitrification. Since the highest NH4-N concentrations occurred in parallel to the methane breakthrough, it appears that oxygen shortage is the dominant cause for incomplete nitrification and methane oxidation. As the methane load exceeded 25 g CH4/(m2⋅d), slight nitrite accumulation occurred in the MET column (0.1 ± 0.05 mg NO2-N/l).

where oPO4-Pconsumed is the consumed oPO4-P load during methane removal, g; and CH4,removed is the removed methane load, g; R2 = 0.80, n = 19. The phosphorous demand for methane removal was calculated as difference of the effluent phosphorous load of both columns. Another indication for biomass washout is the strong correlation of TKN and TSS in the effluent:

CTKN = 0.36·CTSS

CTKN = 5.30·(CNH4 - N )0.47

(10)

where CTKN is the TKN concentration of the effluent, mg/L; and CNH4 - N is the ammonia concentration of the effluent, mg/L; R2 = 0.93, n = 6. It is likely that the observed COD washout results from biomass growth or EPS production. However, only a small part of the biomass was washed out and detected as COD accumulation. The major part of the biomass was retained in the filter. The COD washout averaged 0.07 ± 0.05 g COD/g CH4,removed, while DOC washout averaged 0.02 ± 0.01 g DOC/g CH4,removed. Under typical conditions for VF wetlands, the COD increase in the effluent will amount for up to 7 mg/L or 20%, respectively (assumptions: hydraulic load: 60–100 L/(p∙d), COD load: 10–20 g/(m2∙d), COD removal efficiency: 95%). Hence, low CH4 loads have only a minor influence on wastewater treatment. Due to the short examination period, no final conclusions about the long-term behavior of combined wastewater treatment and methane oxidation are drawn. However, it is clear that negative impacts of methane removal on COD washout are minimized at low methane loads.

1.8 CH4,rem. = 15.5 · COD acc. + 3.3 R² = 0.87

25

1.5

20

1.2

15

0.9 NO3-Nden. = 1.40 · COD acc. - 0.68 R² = 0.87

10

rem. CH4/acc. COD CH4/COD den. NO3-N/acc. COD NO3-N/COD

5

0.6 0.3

0

3.2.7. COD balancing For COD balancing two approaches were used (see section ‘methods’). Due to the different initial values, a divergence was detected (Eq. (11)), which increases with rising methane load.

Denitrified nitrate [g/(m² d)]

Removed CH4 [g/(m² d)]

(9)

where CTKN is the TKN effluent concentration of the MET column, mg/ L; and CTSS is the effluent concentration of Total Suspended Solids of the MET column, mg/L; R2 = 0.97, n = 6. In contrast to NH4-N, TKN was analyzed only from the last six effluent samples of the MET column. The ratio between both parameters was used to calculate the missing TKN concentrations of the MET column:

3.2.6. COD accumulation and biomass growth In parallel to methane oxidation, COD accumulation was observed in column MET (Fig. 5). Accumulation was calculated as difference between the COD effluent load of both columns. The COD removal efficiency dropped from 95% without methane feed to 88% with methane feed while the efficiency remained constant at 95% in the reference column. Prior to methane feeding, the effluent COD concentrations averaged 14.4 ± 4.5 mg/L in the MET column and 14.3 ± 4.2 mg/L in the reference column. Before methane breakthrough occurred, the effluent COD concentrations increased in the MET column with methane feeding on average to 17.6 ± 8.5 mg/L, while effluent concentrations in the reference column averaged 12.8 ± 2.7 mg/L. In the last six weeks of operation, effluent COD concentrations in column MET increased further to 68 ± 20 mg/L with a peak value of 100 mg/L (Fig. 5). In the reference column, the corresponding concentrations averaged 15.0 ± 4.5 mg/L (Fig. 5). Accumulated COD and removed methane correlate (Fig. 6), which implies a direct impact of methane feeding to effluent COD. Since the correlation between accumulated COD and denitrified nitrate is based only on values gathered at methane loads > 9.5 g CH4/(m2∙d) (see Section 3.2.4), the shown trend is present but less substantiated (Fig. 6). It appears that accumulated COD is not an intermediate originating during methane oxidation. No significant concentrations of formic acid, acetic acid, citric acid, hydroxybutyric acid or formaldehyde were 30

(8)

CODdivergence = CODMOM,app1 − CODMOM,app2

0.0 0.0

0.5 1.0 1.5 Accumulated dissolved COD [g/(m² d)]

(11)

where CODMOM,app1 and CODMOM,app2 is the COD of the biomass of the methane oxidizing bacteria calculated according approach 1 and approach 2, g; and CODdivergence is the difference between both approaches, g. Biomass growth associated with MOM resulting from the second approach is always smaller than that resulting from the first approach. Results for the COD divergence are in the range of 12–1651 mg COD or 0.5–8.4%. The values correlate with the COD effluent load of the MET

2.0

Fig. 6. Impact of the removed CH4 on the accumulated dissolved COD, and ratio between the denitrified nitrate and the accumulated dissolved COD. 84

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COD effluent load - COD divergence COD effluent load - fitted COD removal efficiency

2.0

80

1.5

60

1.0

40

0.5

20

0.0

0 0.0

0.5

1.0 1.5 COD effluent load [g]

2.0

load (methane and wastewater), g/(m2⋅d); R2 = 0.80. The estimated CH4 loss was not confirmed analytically. If methane was dissolved in the effluent, it influenced the COD analysis. Due to the height difference between the column outlet and the effluent tank as well as changes of the partial pressure of methane in the sand pores compared to the air outside the column, degassing effects are possible. This would result in reduced methane concentrations at the effluent tank in comparison to the column.

100 Fitted COD removal efficiency [%]

COD divergence [g]

2.5

3.3. Performance assessment

2.5

There are no reference values for constructed wetlands, which can be used to assess the results. But a similar technique is utilized for reducing methane emissions of landfills. Methane oxidation in landfill cover soils is described by several authors (see Kightley et al., 1995; Gebert et al., 2011; Rachor et al., 2011; Roncato and Cabral, 2012; Ndanga et al., 2015). The methane load of landfill cover soils is higher compared to the load of constructed wetlands presented in the actual study. Roncato and Cabral (2012) carried out field tests with loads between 20 and 580 g CH4/(m2⋅d). The maximum oxidation rate was 576 g CH4/(m2⋅d) while the maximum oxidation rate during laboratory trials was lower and reached approximately 120 g CH4/(m2⋅d). Results of further studies show similar and different oxidation rates compared to the results given by Roncato and Cabral (2012). Kightley et al. (1995) achieved oxidation rates between 109 and 166 g CH4/(m2⋅d), Rachor et al. (2011) achieved 39–80 g CH4/(m2⋅d). Oxidation rates published by Ndanga et al. (2015) were between 105 and 253 g CH4/(m2⋅d). The published data show higher oxidation rates compared to the presented study. The differences between the methane oxidation in landfill covers and the described technique are mainly due to the different test set-up and the different treatment goals. On the one hand, the CW was fed discontinuously with methane, on the other hand, the filter layer height was lower than during the studies dealing with landfill covers. In literature, the oxidation layer height above the methane inlet was approximately 0.8–1.0 m (see Kightley et al., 1995; Gebert et al., 2011; Rachor et al., 2011; Roncato and Cabral, 2012; Ndanga et al., 2015), while in the presented study, the filter height above the methane inlet was 0.3 m. Therewith, methane oxidation in landfill covers is backed by larger treatment volumes. During methane oxidation in landfill covers, sand-compost-mixtures (Roncato and Cabral, 2012), sand-gravel-mixtures and sand (Kightley et al., 1995; Rachor et al., 2011; Ndanga et al., 2015) combined with topsoils (Ndanga et al., 2015) are utilized. Hence, the chosen approach utilizing sand of a CW is suitable. According to Gebert et al. (2011), sand is suited due to its air filled porosity, which is important for gas transport and oxidation (Rachor et al., 2011). Additionally, a higher porosity reduces negative impacts of formed EPS, such as clogging of the filter (Streese and Stegmann, 2003). In consideration of the low methane load generated in pretreatment units, it is expected that clogging due to EPS has a minor impact on operation of CWs. In contrast, EPS washout can have an impact on COD effluent concentration. This fact does not have importance in case of treating methane emissions of landfills, but in case of CWs the threshold values must be met. Compared with landfill covers, both wastewater treatment and complete methane removal are on priority for CWs. Since the highest methanotrophic activity occurs within half a meter below the surface (Chiemchaisri et al., 2001), the tested configuration supplies proper conditions for oxidation of COD, NH4-N and CH4.

Fig. 7. Relation of the COD effluent load to the COD divergence between both balancing approaches and the fitted COD removal.

column (Fig. 7). It appears that methane loading has a negative impact on COD removal, possibly caused by oxygen shortage or biomass washout. Since the divergence is included in the COD removal efficiency according to Eq. (12) (‘fitted efficiency’), a correlation between the fitted efficiency and the COD effluent load is observed (Fig. 7).

CODdivergence ⎞ CODREF,ww,eff. CODMET,ww,rem. = ⎜⎛1 − − ⎟ ·100% CODREF,ww,feed CODMET,ww,feed ⎠ ⎝

(12)

where CODMET,ww,rem. is the removal efficiency for the wastewater COD of the MET column, %; CODREF,ww,eff. is the COD effluent load of the REF column, g; CODREF,ww,feed. is the COD inflow load of the REF column, g; CODdivergence is calculated according Eq. (11), g; and CODMET,ww,feed is the COD inflow load of the wastewater fed to the MET column, g. Results of balancing are shown for three operating conditions in Fig. 8 including the fitted efficiency for the removal of the COD contained in wastewater. It confirms that the major part of the removed methane was converted to biomass or biomass byproducts (see Section 3.2.6). Accordingly, the efficiency of the COD removal of the fed wastewater was observed to drop to about 80%, when the CH4 surface load did not exceed 20 g CH4/(m2∙d). Higher methane loads caused a distinct increase of COD effluent concentrations. From the collected data, it is impossible to identify whether the cause for this is biomass washout or an impaired degradation of the fed wastewater substances. The total removal efficiency, relating to the fed wastewater COD, the fed methane and the measured COD concentration in the effluent averaged 97 ± 1.7%. Lower values corresponded to high load conditions with methane breakthrough. Including the calculated washout of dissolved CH4 according to Eq. (13) (see also Section 3.2.4 and Fig. 8), the removal efficiency averages 88 ± 5.8% (72–93%):

CODCH4,HS + CODCH4 ,dis. + CODww,eff. ⎞ CODMET,rem. = ⎜⎛1 − ⎟ ·100% CODCH4 ,feed + CODww,feed ⎝ ⎠ (13) where CODMET,rem. is the total efficiency for COD removal of the MET column (methane and wastewater), %; CODCH4,HS is the COD equivalent load of the CH4 contained in the headspace, g; CODCH4,dis. is the calculated load of the dissolved CH4, g; CODww,eff. is the COD load of the effluent, g; CODCH4,feed is the COD equivalent load of the fed CH4, g; and CODww,feed is the COD load of the fed wastewater, g. The efficiency increases with rising COD surface load due to the higher impact of CH4 washout during low COD surface loads:

CODMET,rem. = 54.4·(CODSL )0.11

4. Conclusions Vertical subsurface flow constructed wetlands are suitable for combined methane oxidation and wastewater treatment. Due to the low methane generation during anaerobic degradation processes in pretreatment units of small sewage works, the methane load does not

(14)

where CODMET,rem. is the total efficiency for COD removal of the MET column (methane and wastewater), %; and CODSL is the COD surface 85

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Fig. 8. COD balancing during different load conditions (column MET).

following changes on the set-up are considered:

substantially impact wastewater treatment processes. The results indicate that the major part of methane was removed from the fed gas but not oxidized. Methane degradation leads to increased biomass production. The more CH4 is oxidized, the more biomass is formed. Since biomass is washed out of the filter, methane oxidation causes an increase of COD effluent concentrations. At high CH4 loads, the COD can increase by up to 90 mg/L of which 90% is caused by washout. During low load conditions, the COD effluent concentrations are affected to a negligible degree. Since oxygen is the major electron acceptor for methane oxidation, methane degradation via nitrate is of minor importance for denitrification. Nevertheless, it is demonstrated by mass balancing that a fraction of denitrification processes is coupled to methane oxidation, although a high CH4 load is necessary for considerable denitrification activity. The process described is suitable to minimize methane emissions from pretreatment units of small VF wetlands. Based on the results the

- Changing the position of the gas inlet system to the bottom of the sand layer for enhancing the methane oxidation volume, - Changing intermittent methane feeding to continuous methane feeding due to the minor effect of methane on denitrification, - Using of coarse sand to minimize clogging due to EPS formation. But to ensure complete nitrification, sand with a maximum grain size of 0–4 mm should be used.

Acknowledgments This work was supported by the German Federal Ministry for Economic Affairs and Energy (grant number KF2488502RH2).

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