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Improving denitrification in an aquaculture wetland using fish waste - a case study
Ecological Engineering 143 (2020) 105686
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Improving denitrification in an aquaculture wetland using fish waste - a case study
T
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Mathis von Ahnen , Per Bovbjerg Pedersen, Johanne Dalsgaard Technical University of Denmark, DTU Aqua, Section for Aquaculture, The North Sea Research Centre, P.O. Box 101, DK-9850 Hirtshals, Denmark
A R T I C LE I N FO
A B S T R A C T
Keywords: Organic matter Nitrate Constructed wetland
Cost-efficient, end-of-pipe, nitrate removal techniques are called for by the commercial aquaculture industry. This case study examined how simple flow manipulations improved the denitrification performance of a 19,007 m2 (13,305 m3) constructed, free water surface (FWS) wetland treating aquaculture effluent. The wetland consisted of two separate streams with a common outlet: one stream treating nitrate-rich but carbon deficient effluent from the production unit at a hydraulic retention time (HRT) of 1.5 days (wetland stream 1); and a second stream treating carbon-rich, fish sludge-based effluent at a HRT of 41.0 days (wetland stream 2). During the course of the study (May–July 2017), three increasing proportions (40, 49 and 56%) of nitrate-rich effluent were re-directed from wetland stream 1 to the sludge-fed wetland stream 2 aiming at improving heterotrophic denitrification conditions in wetland stream 2 and consequently nitrogen removal in the wetland as a whole. Inlet C/N ratio in wetland stream 2 decreased from 1086 ± 57 to an average of 234 ± 56 (p < .05), and the area-based, total nitrogen (TN) removal rate in this wetland section increased significantly from 0.1 ± 0.01 to 8.4 ± 1.4 g/m2/d at the highest manipulated flow. In comparison, the flow manipulations had no effect on TN removal rates in wetland stream 1 averaging 1.4 ± 0.2 g/m2/d throughout the study. For the wetland as a whole, the TN removal rate increased from 1.4 ± 0.2 to 3.9 ± 0.8 g TN/m2/d. The flow manipulations furthermore improved the removal rates of total phosphorous and dissolved organic matter in the wetland as a whole. The study demonstrates that denitrification in a constructed aquaculture wetland may be improved by combining sludge-based and nitrogen-rich effluents in right proportions and leading it through an anoxic section of the wetland.
1. Introduction
horizontal-flow constructed wetlands treating aquaculture effluent (e.g., Schulz et al., 2003; Schulz et al., 2004; Sindilariu et al., 2007; Sindilariu et al., 2008; Sindilariu et al., 2009; Dalsgaard et al., 2018). Schulz et al. (2004) investigated the effect of hydraulic retention time (HRT) in three parallel, 350 m2 free water surface (FWS) wetlands treating effluent from a flow-through rainbow trout (Oncorhynchus mykiss) farm. Mean total nitrogen removal rates of 0.45, 0.71, and 0.82 g/m2/d at HRTs of 3.5, 5.5, and 11 h, respectively, were found, and the authors concluded that a shorter HRT favored aerobic conditions and hindered denitrification. Dalsgaard et al. (2018) investigated longitudinal and seasonal nutrient removal rates in a FWS wetland treating effluent from a Danish Model Trout Farm (MTF) type I (Jokumsen and Svendsen, 2010). The wetland was characterized by a relatively short retention time (7.7 h), and a miniscule net removal of nitrate-N was observed in spring (February–June) with area-based removal rate constants (kA) averaging 0.1 ± 0.1 m/d. This was succeeded by a small net production of nitrate-N from July to January, and
Cost-efficient, end-of-pipe techniques for removing nitrate from recirculating aquaculture system (RAS) effluent are required to reduce nitrogen discharge from aquaculture farms. Constructed wetlands are man-made treatment systems successfully applied in many parts of the world for treating municipal, agricultural, and industrial wastewater (Verhoeven and Meuleman, 1999). They are mechanically simple but biologically complex systems. Nutrient / pollutant removal happens in a rather “uncontrolled” manner based on biological and biochemical processes affected by interactions between plants, microorganisms, and the sediment (Kadlec and Knight, 1996). Where space allows, constructed wetlands may also be established for treating dilute aquaculture effluent. To improve effectiveness of such engineered systems, underlying factors governing nutrient removal processes need to be understood and ultimately controlled. Several studies have investigated the efficiency of subsurface- and
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Corresponding author. E-mail address: [email protected] (M. von Ahnen).
https://doi.org/10.1016/j.ecoleng.2019.105686 Received 25 June 2019; Received in revised form 8 November 2019; Accepted 20 November 2019 0925-8574/ © 2019 Elsevier B.V. All rights reserved.
Ecological Engineering 143 (2020) 105686
M. von Ahnen, et al.
termed “wetland stream 1” and “wetland stream 2” (Fig. 1 and Table 1). Wetland stream 1 had a total FWS area of 7359 m2. It was formed as a long, narrow channel and received the production unit overflow (3474–3737 m3/d) entering at station A (Fig. 1). In addition, it received the overflow (43 m3/d, station B in Fig. 1) from small holding tanks storing drum filter reject water before pumped to the sludge tank. Wetland stream 1 was operated at a HRT of 1.5 days prior to this study (Table 2). Wetland stream 2 had a total FWS area of 11,078 m2. Water firstly entered two, interconnected basins succeeded by an alternating straight and meandering channel. Shortly before the final wetland outlet, water from wetland stream 2 mixed with water from wetland stream1 (Fig. 1, station L). Wetland stream 2 received sludge pond overflow on Sundays (314 m3; Fig. 1, station G) and biofilter backwash (144 m3/d; Fig. 1, station F) on Mondays – Saturdays (see section 2.1.1). It was operated at a hydraulic retention time of 41.0 days prior to this study calculated by averaging out discrete sludge-based input over a week.
it was argued that denitrification was limited by high oxygen concentrations and low organic carbon availability. In comparison, a two-year monitoring study in eight constructed wetlands associated with Danish MTFs type III found average removal rates of 1.5–1.9 g NO3-N/m2/d (Svendsen et al., 2008). All eight wetlands were characterized by relatively long HRTs (20–50 h), and farm effluent concentrations were higher in organic carbon and nitrate-N than in the wetland study by Dalsgaard et al. (2018) (approximately 10.0 versus 3.5 mg total BOD5/l, and 6.6 versus 3.4 mg NO3-N/l). These average removal rates are, however, still relatively low compared to other denitrification technologies applied in aquaculture (van Rijn et al., 2006) and the current, more or less passive, operation of aquaculture wetlands prevents many Danish farmers from increasing their production because of exceeding their discharge allowance (Danish Ministry of the Environment, 2012). A study by Suhr et al. (2013) demonstrated that RAS sludge may be used as a cost-free carbon source in a single-sludge denitrification process aimed at removing nitrogen from fish farm effluent. By mixing nitrate-rich effluent from a production unit with carbon-rich sludge in a heterotrophic denitrification reactor, nitrogen was removed at a maximum rate of 124.8 ± 15.7 g NO3-N/m3 reactor/d. The study highlighted that C/N ratios and HRT are essential parameters with respect to maximizing nitrogen removal in this type of setup. As an alternative (or in addition) to using a separate, rather complex filter set-up as the one by Suhr et al. (2013), we hypothesized that the sludge denitrification process may be exploited within the basins of an existing constructed wetland. To examine this, an 11-week field study was carried out with the aim of improving overall nitrogen removal in a constructed FWS wetland treating effluent from a Danish Model Trout Farm type III. The objective of the study was to mix internally generated, carbon-rich sludge effluent with increasing proportions of nitraterich effluent in a wetland side stream and measure the effect on overall nitrogen removal rates.
2.2. Experimental protocol The study was divided into two phases: 1) a “phase I" in which the original (baseline) nitrogen removal performance of the wetland was documented; and 2) a flow manipulation phase (“phase II") in which increasing proportions of nitrate-rich inflow to wetland stream 1 were re-directed to the sludge-fed wetland stream 2, and effects on wetland nitrogen removal performance evaluated. 2.2.1. Phase I–Baseline sampling Phase I documented the performance of the two wetland streams and the constructed wetland as a whole (CW) before any flow manipulations. Five, 24 h pooled sampling stations (stations A, B, C, D, and E in Fig. 1) were positioned along wetland stream 1 to follow the longitudinal turnover of nutrients and organic matter. In wetland stream 2, two 24 h pooled sampling stations (H and J in Fig. 1) were established within the wetland. In addition, to two grab sampling stations were established; one for biofilter backwash sampling (station F in Fig. 1) and one for sampling the weekly sludge pond discharge (station G in Fig. 1). Finally, a 24 h pooled sampling station was placed at the final wetland outlet (station L in Fig. 1). Baseline samples were obtained once a week for three consecutive weeks (May 10, 17, and 24, 2017; i.e., n = 3). Dissolved oxygen, pH, and temperature were measured on sample collection. Each day, 3100 kg of feed were administered to the RAS during sampling phase I.
2. Materials and methods 2.1. Study site 2.1.1. Fish farm and operation The study was carried out at a Danish MTF type III producing approximately 600 t rainbow trout/y. The inlet water to the farm consisted of ~25 l/s of groundwater and ~25 l/s of drainage water, which was recovered from below the associated constructed wetland and equally distributed into six independent RAS units. Each unit contained two parallel raceways, sludge cones, drum filters, and a fixed bed biofilter, and recirculation flows were generated via airlifts. Each biofilter was separated into six sections (24 m3 each section), each of which was backwashed on different weekdays, leaving one day per week without backwashing. Backwashing was carried out by disrupting the water flow to a section, injecting air from the bottom for about 10 min, and discharging the water from the section into an adjacent, constructed FWS wetland (Fig. 1). Sludge cones were emptied once a day, and collected sludge and drum filter backwash was pumped into a concrete sludge holding tank (r = 10 m, h = 3.5 m, V = 1100 m3). Once a week, the top 1 m dissolved layer (V = 314 m3) was discharged to a sludge pond (V = 100 m3) that fed into the adjacent wetland. The holding tank was completely emptied twice a year and the sludge transported away.
2.2.2. Phase II–Flow manipulations In study phase II, increasing proportions of nitrate-rich effluent from the production unit were re-directed at station A to the sludge-fed wetland stream 2, aiming at improving overall nitrogen removal in this wetland side stream and in the wetland as a whole. At the same time, the number of 24 h pooled sampling stations in wetland stream 2 was increased from two to five (Fig. 1 stations A, H, I, J, K) while it was reduced from five to three in wetland stream 1 (Fig. 1, stations A, B, E). Flows were manipulated on June 7, June 29 and July 13, 2017. This was done by the farm manager, increasing the overflow volume (via regulation of weir height) from station A into wetland stream 2 from 0 to 1431, 1773, and 2015 m3/d, respectively. The redirected flows corresponded to approximately 40, 49, and 54% of the former wetland stream 1 inflow, and are henceforth referred to as flow 1, 2, and 3. The aim was to establish three increasing flows up to the maximum possible gravity flow in wetland stream 2, without manipulating the weir that delivered water to wetland stream 1. Flow manipulations were succeeded by acclimatization periods of 20, 12, and 12 days, respectively, prior to sampling, corresponding to 5.3, 4.0, and 4.5 times the retention time in wetland stream 2. Pooled 24 h samples and grab samples were obtained for three consecutive days (i.e., n = 3) following each acclimatization period. The farm fed
2.1.2. Constructed wetland A constructed wetland with a total FWS area of 19,007 m2 and an average depth of 0.7 m was associated with the farm for effluent polishing before final discharge to a local stream. The wetland had been in use for 10 years prior to the study, and was constructed by interconnecting former earthen fish ponds (emptied for fish and no longer in used for fish farming) into two, parallel wetland streams henceforth 2
Ecological Engineering 143 (2020) 105686
M. von Ahnen, et al.
Figure 1. Overview of the constructed free water surface wetland treating the effluent from a recirculating Model Trout Farm type III. Indicated is the monitored stretches of the two wetland streams referred to as wetland stream 1 (blue) and wetland stream 2 (red), direction of water flow, and sampling stations (A-L). Orange lines indicate non-monitored stretches prior to the shared outlet (final discharge from the entire constructed wetland) at monitoring station L. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
using automatic, refrigerated samplers (Glacier® Portable, Teledyne ISCO, USA) programmed to sample 300 ml with 1 h intervals for 24 h. Subsamples were filtered through 0.45 μm mixed cellulose ester filters (Whatman, GE Healthcare, UK) prior to analyses of: dissolved five-day biochemical oxygen demand (BOD5-diss; ISO 5815-2, 2003); dissolved chemical oxygen demand (CODdiss.; ISO 15705, 2002; 4–40 mg/l, 15–150 mg/l, 150–1000 mg/l Spectroquant® test kits, Merck Millipore, Germany); and orthophosphate (PO4-P). Other subsamples were filtered through 0.2 μm syringe filters (Filtropur S 0.2 μm, Sarstedt, Germany) prior to analyses of: total ammonia nitrogen (TAN; DS 224, 1975); nitrite-N (NO2-N; DS 223, 1991); nitrate-N (NO3-N; ISO 7890–1, 1986); and dissolved total N (TNdiss; ISO 7890–1, 1986; ISO 11905–1, 1997). Finally, unfiltered sub-samples were analyzed for total BOD5 (BOD5-total; ISO 5815-2, 2003); total COD (CODtotal; ISO 15705, 2002; 4–40 mg/l, 15–150 mg/l, 150–1000 mg/l Spectroquant® test kits, Merck Millipore, Germany); total N (TN; ISO 7890–1, 1986; ISO 119051, 1997); and total phosphorous (TP; ISO 6878, 2004). Samples were kept cold (3 °C) until analysis. Analyses of BOD5 were initiated on the day of sample collection while all other analyses were performed within the first few days after collection. Unfiltered and filtered subsamples for CODtotal, CODdiss, PO4-P, and TP analyses were conserved until analysis by adding 1% sulfuric acid (H2SO4; 21.4%; Merck KGaA, Germany). Dissolved oxygen (mg O2/l), pH, and temperature (°C) were measured at each sampling station using Hach Lange HQ40 multimeters (Hach Lange, Germany).
Table 1 Wetland sampling stations cf. Fig. 1 positioned at increasing, free water surface (FWS) distance from the inlet. Wetland section
Sampling station
FWS area (m2)
Wetland stream 1
A (production unit effluent) B (drum filter reject water) C D E F (biofilter backwash) G (sludge overflow) H I J K L (wetland outlet)
0 0 2934 4625 7359 0 0 2512 5138 6293 8344 19,007
Wetland stream 2
Entire wetland (CW)
2800, 2600, and 3100 kg feed per day, respectively, during the three sampling periods. Flow rates into wetland stream 1 and 2 were derived each sampling day by measuring weir overflow heights and applying Torricelli's law (Molitor, 1908). Additional inflow contributions at station B were measured at an open pipe using a bucket and a stop watch, while inflows at stations F and G were calculated (geometrically) from known tank volumes discharged into the wetland. Wetland outflow rates were assumed to equal inflow rates, i.e., infiltration, precipitation, evapotranspiration, etc. in the wetland was assumed to be negligible. 2.3. Sample analyses
2.4. Calculation of removal rates and statistical analyses Area-based removal rates (g/m2/d) were calculated as the mass
Pooled samples were collected from the centre of the water column
Table 2 Flow rates, hydraulic retention times (HRTs), and influent C/N ratios (as CODdiss/NO3-N) measured during sampling phase I and II. Values in same columns not sharing a common superscript are significantly different (P < .05). Wetland stream 1
Wetland stream 2
Sampling phase
Inflow st. A (m3/ d)
Inflow st. B (m3/ d)
HRT (d)
Influent C/N ratio
Inflow st. A (m3/ d)
Inflow st. F⁎ (m3/ d)
Inflow st. G⁎ (m3/ d)
HRT (d)
Influent C/N ratio
Phase Phase Phase Phase
3474 2146 1860 1722
43 43 43 43
1.5 2.4 2.7 2.9
1.4 1.3 1.1 1.1
0 1431 1773 2015
144 144 144 144
45 45 45 45
23.3 2.7 2.3 2.0
1086 ± 57a 216 ± 54b 233 ± 55b 261 ± 49b
⁎
I, baseline II, flow 1 II, flow 2 II, flow 3
± ± ± ±
0.3 0.1 0.2 0.2
Calculated by averaging out discrete sludge-based input over a week. 3
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Table 3 Removal rates in g/m2/d⁎ (mean ± SD, n = 3, for BOD5n = 2) of various nutrients for wetland stream 1 (W1), wetland stream 2 (W2)⁎⁎, and the entire constructed wetland (CW) as well as average wetland temperatures (°C) obtained during measurements of the baseline performance (phase I) and the flow manipulations (flows 1–3, phase II). Nutrient
Wetland
Baseline
Flow 1
Flow 2
Flow 3
TN
W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW CW
1.4 ± 0.4 0.1 ± 0.0a 1.4 ± 0.2a 1.2 ± 0.6 1.7 ± 0.1a 1.4 ± 0.2a 1.1 ± 0.2 0.3 ± 0.1a 0.3 ± 0.6a 0.13 ± 0.05 0.02 ± 0.01 0.08 ± 0.03 0.2 ± 0.2 1.6 ± 0.2a 0.5 ± 0.2 0.02 ± 0.01ab 1.00 ± 0.02a 0.27 ± 0.08a 0.03 ± 0.03 0.62 ± 0.05a 0.15 ± 0.06a 0.4 ± 2.6 45.1 ± 1.4 16.0 ± 1.0 2.0 ± 0.5 30.5 ± 0.9a 10.7 ± 0.2a 1.0 ± 1.2 21.7 ± 14.1 7.2 ± 4.2 1.4 ± 0.3 14.1 ± 9.3 5.3 ± 3.2a 13.3 ± 1.8 a
1.1 ± 0.4 4.8 ± 0.3b 2.2 ± 0.2b 1.1 ± 0.3 2.3 ± 0.2ab 1.3 ± 0.2ab 0.8 ± 0.3 3.1 ± 0.4b 1.4 ± 0.1b 0.05 ± 0.04 −0.12 ± 0.13 0.01 ± 0.01 0.1 ± 0.1 1.8 ± 0.1a 0.2 ± 0.6 0.01 ± 0.02a 1.42 ± 0.03b 0.54 ± 0.03b −0.01 ± 0.01 1.27 ± 0.11b 0.46 ± 0.06b 2.3 ± 0.6 91.7 ± 34.5 32.4 ± 12.3 1.7 ± 0.3 59.3 ± 1.4ab 20.6 ± 0.9b 2.2 ± 0.4 35.1 ± 4.9 16.7 ± 2.4 1.3 ± 0.1 34.4 ± 1.7 15.8 ± 0.9b 14.8 ± 0.8 b
1.4 ± 0.1 5.5 ± 0.1c 2.8 ± 0.1c 1.4 ± 0.2 2.2 ± 0.1ab 1.7 ± 0.1ab 1.0 ± 0.1 3.6 ± 0.3b 2.1 ± 0.1b 0.01 ± 0.02 −0.08 ± 0.02 0.02 ± 0.01 0.1 ± 0.0 1.9 ± 0.2a 0.6 ± 0.0 0.04 ± 0.01ab 1.49 ± 0.11c 0.43 ± 0.02ab 0.01 ± 0.01 1.27 ± 0.12c 0.38 ± 0.04ab 1.8 ± 0.6 73.0 ± 0.4 25.0 ± 0.4 1.4 ± 0.5 64.2 ± 1.3ab 21.9 ± 0.6c 1.3 ± 0.2 40.5 ± 1.8 19.4 ± 0.2 0.7 ± 0.3 39.2 ± 0.6 17.9 ± 0.8b 15.4 ± 0.7 c
1.8 ± 1.0 8.4 ± 1.4d 3.9 ± 0.8c 1.7 ± 1.1 7.4 ± 1.1b 3.7 ± 0.9b 1.2 ± 1.1 5.1 ± 1.7b 2.7 ± 1.1b 0.03 ± 0.04 0.00 ± 0.05 0.04 ± 0.02 0.2 ± 0.2 3.4 ± 0.3b 1.1 ± 0.3 0.06 ± 0.01b 1.99 ± 0.05d 0.56 ± 0.08b 0.02 ± 0.02 1.88 ± 0.33d 0.54 ± 0.16b 2.3 ± 0.5 98.9 ± 1.1 33.2 ± 0.3 1.5 ± 0.3 86.8 ± 6.8b 29.3 ± 2.4d 1.7 ± 0.4 43.9 ± 1.6 20.4 ± 0.5 0.9 ± 0.4 43.6 ± 0.9 19.8 ± 0.7b 16.0 ± 0.7 c
TNdiss
NO3-N
NO2-N
TAN
TP
PO4-P
CODtotal
CODdiss
BOD5-total
BOD5-diss
Temperature
⁎ Values in the same row not sharing a common superscript letter are significantly different (P < .05). The absence of letters in rows indicates that values are not significantly different. ⁎⁎ Note that removal rates for wetland 2 are based on the area between the wetland inlet and station J as station K was not sampled during baseline sampling (section 2.1.2).
comparison, there was no change in total nitrogen removal in wetland stream 1, ranging between 1.1 ± 0.4 and 1.8 ± 1.0 g TN/m2/d throughout the study (Table 3). These results sustain that denitrification, prior to flow manipulations, was nitrate-limited in wetland stream 2, while it was limited by low carbon availability and high oxygen concentrations in wetland stream 1 (Fig. 2e) as previously observed in an aquaculture wetland operated at short HRT (Dalsgaard et al., 2018). Hence, nitrate limitation typically occurs at NO3-N concentrations below 1 mg/l (Henze et al., 2002). Consistent with this, an average NO3-N concentration of 0.6 ± 0.3 mg/l was measured at station J in wetland stream 2 during baseline sampling. At such low NO3-N concentrations, available organic matter typically fuels heterotrophic bacterial processes such as sulfate reduction and methanogenesis rather than heterotrophic denitrification (Korom, 1992). During flow manipulations, the nitrate load on wetland stream 2 increased from a baseline value of 0.3 ± 0.2 to 4.6 ± 0.3, 5.8 ± 0.1, and 7.4 ± 1.6 g NO3-N/m2/d, respectively, at flow 1, 2 and 3. Corresponding NO3-N concentrations at station J increased from a baseline value of 0.6 ± 0.3 to 5.6 ± 0.4, 6.8 ± 0.8, and 6.7 ± 0.5 mg/l. I.e., the flow manipulations apparently improved the conditions for heterotrophic denitrification in wetland stream 2 to an extent that it was no longer nitrate limited. According to Fig. 2a, most of the nitrogen removal took place prior to station H, inferred from the decline in TN removal rates following this station. Organic matter removal followed a similar pattern (Fig. 2 c,
removed between the wetland inlet station and a specific wetland station (assuming plug flow conditions) and dividing with the FWS wetland area between the two stations. Results are throughout shown as mean ± SD. Statistical analyses were carried out in SigmaPlot version 13.0 (Systat Software Inc., CA, USA) applying a significance level of p < .05. Removal rates following flow manipulations were compared using One-way ANOVA followed by Holm Šidák multiple comparison test in case of significant differences. In case the normality test (Shapiro-Wilk) failed, one-way ANOVA on ranks was run followed by a Tukey multiple comparison test. Spearman correlation analysis was applied for testing the correlation between area-based NO3-N and CODdiss removal rates.
3. Results and discussion 3.1. Nitrogen removal Increasing the inflow of nitrate-rich effluent to wetland stream 2 led to a significant increase in overall nitrogen removal of the entire wetland (CW; Tables 3 and 4). The area-based TN removal rate for the whole wetland increased from a baseline value of 1.4 ± 0.2 to 3.9 ± 0.8 g/m2/d at the highest manipulated flow (flow 3). This improvement was principally due to a significant increase in wetland stream 2 nitrate removal (in area delimited by station J), increasing from a baseline value of 0.3 ± 0.1 to 5.1 ± 1.7 g NO3-N/m2/d. In 4
Ecological Engineering 143 (2020) 105686
55 55 60 31 55 60 65 86 89 93 97 15.5 ± 0.4 14.4 ± 0.9 8.7 ± 1.0 0.43 ± 0.01 4.37 ± 0.55 1.80 ± 0.41 1.37 ± 0.28 26.0 ± 3.5 18.1 ± 0.3 7.0 ± 2.7 2.6 ± 0.2
d), and there was a positive correlation (r = 0.66, p < .001, n = 36) between NO3-N and CODdiss removal rates in the flow manipulation phase. It thus seems that heterotrophic denitrification switched from being nitrate limited prior to flow manipulations to possibly becoming carbon limited downstream wetland stream 2. 3.2. Organic matter removal Organic matter removal rates of both COD and BOD5 were at all times considerably higher in wetland stream 2 than in wetland stream 1 (Table 3), reflecting different hydraulic conditions and organic matter loadings. Flow manipulations led to a significant increase in CODdiss removal rates in wetland stream 2, increasing from a baseline value of 30.5 to 86.8 g/m2/d at the highest manipulated flow. The removal rate of BOD5-diss similarly increased (from 14.1 to 43.6 g/m2/d; Table 3), however, the increase was not statistically significant presumably due to the low number of replicates (n = 2). These improvements in dissolved organic matter removal rates are consistent with the higher denitrification activity observed in wetland stream 2 as discussed in section 3.1. In addition to carbon from added sludge overflow, additional carbon may derive from wetland-internally generated, dead and decomposing plant biomass. However, not all of this carbon is readily available for denitrifiers, and previous wetland studies have shown that carbon can be limiting at high nitrate loading, in which case denitrification will depend strongly on the carbon supply rate and availability (reviewed by Kadlec and Wallace, 2009; Hang et al., 2016). There were no significant changes in organic matter removal rates in wetland stream 1 following flow manipulations. However, total organic matter removal rates were in all cases slightly higher and dissolved organic removal rates slightly lower than at baseline (Table 3). Flow manipulations increased the hydraulic retention time in wetland stream 1 from 1.5 to 2.9 days (Table 2). The time available for breakdown of particulate biodegradable organic matter thereby increased, and this may explain the tendency for a higher area-based removal rate. Dissolved organic matter, on the other hand, was probably fully degraded in wetland stream 1 even before flow manipulations given the high oxygen concentrations (Fig. 2e), and a longer retention time therefore actually reduced the area-based removal rate of dissolved organic matter in wetland stream 1. 3.3. C/N ratios
38 28 37 8 15 72 72 88 86 95 99 28.6 ± 2.3 23.0 ± 1.4 18.6 ± 1.3 0.70 ± 0.08 6.78 ± 0.19 3.75 ± 0.10 3.18 ± 0.24 184.0 ± 60.9 120.0 ± 4.8 87.3 ± 12.1 80.2 ± 4.9
17.6 ± 1.9 16.6 ± 1.0 11.6 ± 1.5 0.64 ± 0.01 5.76 ± 2.72 1.07 ± 0.06 0.88 ± 0.04 22.2 ± 1.5 17.4 ± 1.1 4.0 ± 0.1 1.2 ± 0.3
29.0 ± 0.5 23.0 ± 0.7 19.1 ± 0.3 0.44 ± 0.01 6.91 ± 0.10 3.84 ± 0.08 3.14 ± 0.13 148.1 ± 3.0 125.0 ± 3.7 100.3 ± 2.0 90.4 ± 4.0
15.2 ± 0.1 14.4 ± 0.2 8.7 ± 0.4 0.35 ± 0.04 3.76 ± 0.06 1.71 ± 0.16 1.28 ± 0.07 26.1 ± 3.4 17.3 ± 0.9 4.8 ± 1.0 2.3 ± 0.3
48 37 54 22 46 55 59 82 86 95 97
34.3 ± 3.5 32.0 ± 3.6 21.7 ± 4.6 0.62 ± 0.12 9.75 ± 0.77 4.47 ± 0.09 3.97 ± 0.48 185.2 ± 2.2 158.6 ± 12.0 104.4 ± 0.1 97.5 ± 3.4
Removal% Outlet mg/l Inlet mg/l Removal % Outlet mg/l Inlet mg/l Removal % Outlet mg/l
27 28 12 46 28 50 42 80 77 88 95
Increasing the nitrate load on wetland stream 2 reduced the incoming C/N ratio (calculated as CODdiss/NO3-N), declining significantly from an average of 1086 ± 57 before flow manipulations to an average of 234 ± 56 following flow manipulations (Table 2). Wetland stream 2 presumably received large amounts of easily biodegradable organic matter including volatile fatty acids (VFAs) deriving from hydrolysis in the sludge tank. Prior to flow manipulations, this organic matter was likely “wasted” to aerobic respiration and subsequently to anaerobic respiration such as sulfate reduction and methanogenesis in the lack of nitrate. Introducing nitrate-rich effluent improved the conditions for denitrification, and denitrification activity was high as long as easily decomposable organic matter was available. Moving downstream in wetland stream 2, the fraction of less biodegradable organic carbon, however, increased as also observed by Puigagut et al. (2008) in a study on organic matter in a horizontal subsurface-flow wetland. Relatively high NO3-N concentrations of 6.35 ± 0.52 mg/l were thus measured at station J by the end of the study despite otherwise favorable denitrification conditions including oxygen concentrations below 0.4 mg/l (Fig. 2e) at a C/N ratio of 3.01.
26.0 ± 0.9 25.0 ± 1.5 13.9 ± 1.7 0.85 ± 0.11 8.28 ± 0.50 2.70 ± 0.02 1.85 ± 0.10 101.1 ± 2.8 71.0 ± 0.3 61.6 ± 2.7 28.7 ± 16.4 TN TNdiss NO3-N NO2-N TAN TP PO4-P CODtotal CODdiss BOD5-total BOD5-diss
19.0 ± 1.7 18.0 ± 1.9 12.2 ± 1.4 0.46 ± 0.04 5.99 ± 0.83 1.34 ± 0.40 1.07 ± 0.32 20.6 ± 2.9 16.3 ± 1.4 7.19 ± 4.3 1.4 ± 0.1
Inlet mg/l Parameter
Oultet mg/l
Removal %
Inlet mg/l
Flow 2 Flow 1 Baseline
Table 4 Averaged inlet and outlet concentrations as well as percentage removals during baseline sampling and flow manipulations for the entire constructed wetland (CW).
Flow 3
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3.4. Comparison with previous wetland denitrification studies The highest TN removal rates in the current study, 3.3 g/m2/d for 5
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the wetland as a whole and 8.4 g TN/m2/d for wetland stream 2, may be compared with the 1.5–1.9 g TN/m2/d measured by Svendsen et al. (2008) in a monitoring study of eight aquaculture wetlands. The eight wetlands were all characterized by HRTs of 20–50 h and average inlet COD/(NO2-N + NO3-N) ratios of 5.5. All discharge from the production units was typically mixed with sludge basin overflow prior to the wetlands, explaining the lower C/N ratio compared to the current study. These different results demonstrate that fish sludge, with little effort, may be used as an internal carbon source to improve denitrification in aquaculture wetlands given the right conditions. This includes optimizing the C/N ratio (5 or 10 to 1; reviewed by Kadlec and Wallace, 2009) using recently hydrolyzed sludge, establishing anaerobic zones, and treating only a side stream of the production unit discharge to sustain that easily biodegradable organic matter is not “wasted” in aerobic respiration. The latter was probably the situation in a FWS wetland treating dilute effluent from a Danish recirculating farm at short retention time (0.32 d; Dalsgaard et al., 2018). No nitrate removal was observed in that study during a whole year of monitoring. Rather, the wetland even became a net nitrate producer at certain times presumably due to a lack of readily available organic carbon combined with high oxygen concentrations. Similarly, Schulz et al. (2004) found comparatively low NO3-N removal rates of 0.06 ± 0.14 g/m2/d at a HRT of 11 h, and a net production of up to 0.27 ± 0.12 g NO3-N/m2/d at a HRT of 3.5 h in FWS wetlands treating highly dilute effluent from a flow-through rainbow trout fingerling farm. On the other hand, when too little production effluent mixes with available sludge effluent in a wetland, nitrate limitation may reduce overall nitrogen removal as was the case in wetland stream 2 during baseline sampling (very high C/N ratio).
majority of excreted phosphorous will be bound in solid waste (Dalsgaard and Pedersen, 2011). Solid waste can be removed by sedimentation (Kadlec and Wallace, 2009), and sedimentation of suspended particles may explain some of the higher removal in wetland stream 2. Sedimentation of particulate P probably also explains the similar removal rates obtained by Schulz et al. (2004), studying removal rates in three parallel FWS aquaculture wetlands treating effluent from a flowthrough farm. In that study, TP removal rates increased from 0.12 to 0.30 g TP/m2/d when loadings increased from 0.23 to 0.70 g TP/m2/d. In addition to particle sedimentation, higher denitrification activity in wetland stream 2 may also have improved phosphorous removal. Certain denitrifiers are known to accumulate orthophosphate in excess of their metabolic requirements leading to a considerable reduction in phosphorous load in the water column (van Rijn et al., 2006). In addition, heterotrophic denitrification activity produces alkalinity (van Rijn et al., 2006). Combined with a higher effluent volume entering wetland stream 2, this presumably contributed to raising outlet pH in wetland stream 2 from a baseline value of 6.8 to above 7.0 (p < .05; Fig. 2f). An increase in pH may cause calcium phosphate to precipitate in denitrifying biofilms in which pH can locally increase above values measured in the water (Arvin and Kristensen, 1982). In comparison to wetland stream 2, wetland stream 1 received low concentrations of primarily dissolved phosphorous. Dissolved phosphorous may be removed by sorption, plant and microbial uptake, and sediment accretion (Kadlec and Wallace, 2009). However, at very low concentrations (approaching a non-zero background concentration) there will be no net removal due to internal cycling, including return flux from decomposition processes, and atmospheric gains and losses (Kadlec and Wallace, 2009). This situation was probably the case in wetland stream 1 where almost no P-removal was observed.
3.5. TAN removal
4. Summary and future prospects
Flow manipulations led to a significant increase in the area-based TAN removal rate in wetland stream 2, increasing from a baseline value of 1.6 to 3.4 g/m2/d at the highest flow (Table 3). No similar effect was observed in wetland stream 1, while for the wetland as a whole there was a non-significant increase from a baseline value of 0.5 to 1.1 g/m2/ d (Table 3). The increase in TAN removal in wetland stream 2 may relate to the reduction in the incoming C/N ratio (Table 2), possibly leading to a reduction in DNRA activity (dissimilatory nitrate reduction to ammonia). This process is conducted by fermentative bacteria in excessive reductive environments (i.e., high C/N; Washbourne et al., 2011) as prevailing in wetland stream 2 especially during baseline sampling. A reduction in the C/N ratio may also have favored an increase in autotrophic anaerobic ammonia oxidation (anammox activity). This anaerobic and autotrophic microbial process converts ammonia and nitrite to dinitrogen, and may have lead to additional removal of TAN and TN (Mulder et al., 1995; Zhu et al., 2011; Suhr et al., 2013).
This case study demonstrated that nitrogen removal in an aquaculture wetland can be improved by using internally generated sludge/ fish waste as a carbon source for denitrification in an anoxic part of the wetland. As a side effect, some additional removal of phosphorous may be achieved via the enhanced denitrification activity. Through flow manipulations, it was demonstrated that the farm was able to significantly reduce the discharge of nitrogen and its overall environmental impact(Table 4). For further optimization, controlled biological sludge hydrolysis prior to the wetland, i.e., solubilization and fermentation of settled organic matter to volatile fatty acids (VFAs), may improve the denitrification process (Letelier-Gordo et al., 2015; Letelier-Gordo et al., 2017). Sludge hydrolysis is carried out by naturally occurring facultative anaerobic bacteria, and happens by itself if particulate organic matter is left to settle. However, solubilized organic matter should be transferred to the wetland after a few days as the anaerobic degradation process otherwise continues into a methanogenic stage and degradable carbon will be “lost” as methane (Letelier-Gordo et al., 2015, LetelierGordo et al., 2017). In this particular case study, wetland stream 1 carried a side stream high in nitrate and low in organic carbon. This combination is ideal for further denitrification treatment in a woodchip bioreactor recently demonstrated as another low-cost, end-of-pipe denitrification technology successfully applied in aquaculture (von Ahnen et al., 2018).
3.6. Phosphorous removal Besides improving nitrogen and organic matter removal, flow manipulations also improved the removal rates of TP and PO4-P in wetland stream 2 increasing significantly from 1.00 ± 0.02 to 1.99 ± 0.05, and from 0.62 ± 0.05 to 1.88 ± 0.33 g/m2/d, respectively (Table 3). No changes were observed in wetland stream 1 apart from an infinitesimal improvement in TP removal. Considering the wetland as a whole, there was a relatively small but significant improvement in TP and PO4-P removal (Table 4). Wetland stream 2 received a higher total phosphorous load than wetland stream 1 (1.13 ± 0.02 vs. 0.39 ± 0.03 g TP/m2/d, baseline sampling) including a relatively larger share of particle bound phosphorous from the sludge/particulate waste. When feeding a commercial diet balanced in phosphorous (as in the current case study), the
Authors contributions Mathis von Ahnen participated in the design of the study, carried out the sampling and drafted the manuscript. Per Bovbjerg Pedersen participated in the design of the study and helped to draft the manuscript. Johanne Dalsgaard participated in the design of the study and coordination and helped to draft the manuscript. All authors read and approved the final manuscript. 7
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Declaration of Competing Interest
ISO 7890-1, 1986. Water Quality—Determination of Nitrate. Part 1: 2.6-Dimethylphenol Spectrometric Method. International Organization for Standardization, Geneva, Switzerland. Jokumsen, A., Svendsen, L.M., 2010. Farming of freshwater rainbow trout in Denmark. In: DTU Aqua Rep 219–2010. Technical University of Denmark, Copenhagen. Kadlec, R.H., Knight, R.L., 1996. Treatment Wetlands. CRC Press, Boca Raton, FL, USA. Kadlec, R.H., Wallace, S.D., 2009. Treatment Wetlands, Second edition. CRC Press, Boca Raton, FL, USA. Korom, S.F., 1992. Natural denitrification in the saturated zone: a review. Water Resour. Res. 28 (6), 1657–1668. Letelier-Gordo, C.O., Dalsgaard, J., Suhr, K.I., Ekman, K.S., Pedersen, P.B., 2015. Reducing the dietary protein:energy (P:E) ratio changes solubilization and fermentation of rainbow trout (Oncorhynchus mykiss) faeces. Aquac. Eng. 66, 22–29. Letelier-Gordo, C.O., Larsen, B.K., Dalsgaard, J., Pedersen, P.B., 2017. The composition of readily available carbon sources produced by fermentation of fish faeces is affected by dietary protein:energy ratios. Aquac. Eng. 77, 27–32. Molitor, D.A., 1908. Hydraulics of Rivers, Weirs and Sluices. Wiley, New York. Mulder, A., van de Graaf, A.A., Robertson, L.A., Kuenen, J.G., 1995. Anaerobic ammonium oxidation discovered in a denitrifying fluidized bed reactor. FEMS Microbiol. Ecol. 16, 177–184. Puigagut, J., Caselles-Osorio, A., Vaello, N., Garcia, J., 2008. Fractionation, biodegradability and particle-size distribution of organic matter in horizontal subsurface-flow constructed wetlands. In: Vymazal, J. (Ed.), Wastewater Treatment, Plant Dynamics and Management in Constructed Wetlands. Springer Science + Business Media B.V. Schulz, C., Gelbrecht, J., Rennert, B., 2003. Treatment of rainbow trout farm effluents in a constructed wetland with emergent plants and subsurface horizontal water flow. Aquaculture 217, 207–221. Schulz, C., Gelbrecht, J., Rennert, B., 2004. Constructed wetlands with free water surface for treatment of aquaculture effluents. J. Appl. Ichthyol. 20, 64–70. Sindilariu, P.D., Schulz, C., Reiter, R., 2007. Treatment of flow-through aquaculture effluents in a constructed wetland. Aquaculture 270, 92–104. Sindilariu, P.D., Wolter, C., Reiter, R., 2008. Constructed wetland as a treatment method for effluents from intensive trout farms. Aquaculture 277, 179–184. Sindilariu, P.D., Brinker, A., Reiter, R., 2009. Factors influencing the efficiency of constructed wetlands used for the treatment of intensive trout farm effluent. Ecol. Eng. 35, 711–722. Suhr, I.S., Pedersen, P.B., Arvin, E., 2013. End-of-pipe denitrification using RAS effluent waste streams: effect of C/N-ratio and hydraulic retention time. Aquac. Eng. 53, 57–64. Svendsen, L.M., Sortkjær, O., Ovesen, N.B., Skriver, J., Larsen, S.E., Bouttrup, S., Pedersen, P.B., Rasmussen, R.S., Dalsgaard, A.J.T., Suhr, K., 2008. Modeldambrug under forsøgsordningen. Faglig slutrapport for måle- og dokumentationsprojekt for modeldambrug (in Danish). In: DTU Aqua Res Rep 193–08. Technical University of Denmark. Van Rijn, J., Tal, Y., Schreier, H.J., 2006. Denitrification in recirculating systems: theory and applications. Aquac. Eng. 34, 364–376. Verhoeven, J.T.A., Meuleman, A.F.M., 1999. Wetlands for wastewater treatment: opportunities and limitations. Ecol. Eng. 12, 5–12. von Ahnen, M., Pedersen, P.B., Dalsgaard, J., 2018. Performance of full-scale woodchip bioreactors treating effluents from commercial RAS. Aquac. Eng. 83, 130–137. Washbourne, I.J., Crenshaw, C.L., Baker, M.A., 2011. Dissimilatory nitrate reduction pathways in an oligotrophic freshwater ecosystem: spatial and temporal trends. Aquat. Microb. Ecol. 65, 55–64. Zhu, G., Wang, S., Feng, X., Fan, G., Jetten, M.S.M., Yin, C., 2011. Anammox bacterial abundance, biodiversity and activity in a constructed wetland. Environ. Sci. Technol. 45, 9951–9958.
The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgements This research was funded by the Ministry of Food, Agriculture and Fisheries of Denmark and by the European Union through the European Maritime and Fisheries Fund (EMFF). We thank the fish farmer for allowing us access to his facilities and the farm manager for helpful assistance during the study. In addition, we thank laboratory technician Ulla Sproegel, Brian Møller, and Dorthe Frandsen (DTU Aqua) for invaluable technical assistance in the laboratories. References Arvin, E., Kristensen, G.H., 1982. Precipitation of calcium phosphate and pH-effects in denitrifying biofilms. Water Sci. Technol. 14 (9–11), 1505–1508. Dalsgaard, J., Pedersen, P.B., 2011. Solid and suspended/dissolved waste (N,P,O) from rainbow trout (Oncorynchus mykiss). Aquaculture 313, 92–99. Dalsgaard, J., von Ahnen, M., Naas, C., Pedersen, P.B., 2018. Nutrient removal in a constructed wetland treating aquaculture effluent at short hydraulic retention time. Aquaculture Environment Interactions 10, 329–343. Danish Ministry of the Environment, 2012. Bekendtgørelse om miljøgodkendelse og samtidig sagsbehandling af ferskvandsdambrug (“in Danish”; Executive order on the environmental approval and handling of freshwater aquaculture systems). Lovtidende A nr. 130. Miljøstyrelsen, Miljøministeriet, Denmark. DS 223, 1991. Water Analysis—Determination of the Sum of Nitrite- and NitrateNitrogen. Danish Standards Foundation, Charlottenlund, Denmark. DS 224, 1975. Water Analysis—Determination of Ammonia-Nitrogen. Danish Standards Foundation, Charlottenlund, Denmark. Hang, Q., Wang, H., Chu, Z., Ye, B., Li, C., Hou, Z., 2016. Application of plant carbon source for denitrification by constructed wetland and bioreactor: review of recent development. Environ. Sci. Pollut. Res. Int. 23, 8260–8274. Henze, M., Harremoës, P., Jes la Coeur, J., Arvin, E., 2002. Wastewater Treatment: Biological and Chemical Processes, 3rd edition. Springer Verlag, Berlin Heidelberg, Germany. ISO 11905-1, 1997. Water Quality—Determination of nitrogen. Part 1: Method Using Oxidative Digestion with Peroxide Sulfate. International Organization for Standardization, Geneva, Switzerland. ISO 15705, 2002. Water Quality—Determination of the Chemical Oxygen Demand Index (ST-COD)—Small-Scale Sealed-Tube Method. International Organization for Standardization, Geneva, Switzerland. ISO 5815-2, 2003. Water Quality—Determination of Biochemical Oxygen Demand after N Days (BODn)—Part 2: Method for Undiluted Samples, ISO 5815-2:2003, Modified. International Organization for Standardization, Geneva, Switzerland. ISO 6878, 2004. Water Quality—Determination of Phosphorus—Ammonium Molybdate Spectrometric Method. International Organization for Standardization, Geneva, Switzerland.
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Improving denitrification in an aquaculture wetland using fish waste - a case study
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Mathis von Ahnen , Per Bovbjerg Pedersen, Johanne Dalsgaard Technical University of Denmark, DTU Aqua, Section for Aquaculture, The North Sea Research Centre, P.O. Box 101, DK-9850 Hirtshals, Denmark
A R T I C LE I N FO
A B S T R A C T
Keywords: Organic matter Nitrate Constructed wetland
Cost-efficient, end-of-pipe, nitrate removal techniques are called for by the commercial aquaculture industry. This case study examined how simple flow manipulations improved the denitrification performance of a 19,007 m2 (13,305 m3) constructed, free water surface (FWS) wetland treating aquaculture effluent. The wetland consisted of two separate streams with a common outlet: one stream treating nitrate-rich but carbon deficient effluent from the production unit at a hydraulic retention time (HRT) of 1.5 days (wetland stream 1); and a second stream treating carbon-rich, fish sludge-based effluent at a HRT of 41.0 days (wetland stream 2). During the course of the study (May–July 2017), three increasing proportions (40, 49 and 56%) of nitrate-rich effluent were re-directed from wetland stream 1 to the sludge-fed wetland stream 2 aiming at improving heterotrophic denitrification conditions in wetland stream 2 and consequently nitrogen removal in the wetland as a whole. Inlet C/N ratio in wetland stream 2 decreased from 1086 ± 57 to an average of 234 ± 56 (p < .05), and the area-based, total nitrogen (TN) removal rate in this wetland section increased significantly from 0.1 ± 0.01 to 8.4 ± 1.4 g/m2/d at the highest manipulated flow. In comparison, the flow manipulations had no effect on TN removal rates in wetland stream 1 averaging 1.4 ± 0.2 g/m2/d throughout the study. For the wetland as a whole, the TN removal rate increased from 1.4 ± 0.2 to 3.9 ± 0.8 g TN/m2/d. The flow manipulations furthermore improved the removal rates of total phosphorous and dissolved organic matter in the wetland as a whole. The study demonstrates that denitrification in a constructed aquaculture wetland may be improved by combining sludge-based and nitrogen-rich effluents in right proportions and leading it through an anoxic section of the wetland.
1. Introduction
horizontal-flow constructed wetlands treating aquaculture effluent (e.g., Schulz et al., 2003; Schulz et al., 2004; Sindilariu et al., 2007; Sindilariu et al., 2008; Sindilariu et al., 2009; Dalsgaard et al., 2018). Schulz et al. (2004) investigated the effect of hydraulic retention time (HRT) in three parallel, 350 m2 free water surface (FWS) wetlands treating effluent from a flow-through rainbow trout (Oncorhynchus mykiss) farm. Mean total nitrogen removal rates of 0.45, 0.71, and 0.82 g/m2/d at HRTs of 3.5, 5.5, and 11 h, respectively, were found, and the authors concluded that a shorter HRT favored aerobic conditions and hindered denitrification. Dalsgaard et al. (2018) investigated longitudinal and seasonal nutrient removal rates in a FWS wetland treating effluent from a Danish Model Trout Farm (MTF) type I (Jokumsen and Svendsen, 2010). The wetland was characterized by a relatively short retention time (7.7 h), and a miniscule net removal of nitrate-N was observed in spring (February–June) with area-based removal rate constants (kA) averaging 0.1 ± 0.1 m/d. This was succeeded by a small net production of nitrate-N from July to January, and
Cost-efficient, end-of-pipe techniques for removing nitrate from recirculating aquaculture system (RAS) effluent are required to reduce nitrogen discharge from aquaculture farms. Constructed wetlands are man-made treatment systems successfully applied in many parts of the world for treating municipal, agricultural, and industrial wastewater (Verhoeven and Meuleman, 1999). They are mechanically simple but biologically complex systems. Nutrient / pollutant removal happens in a rather “uncontrolled” manner based on biological and biochemical processes affected by interactions between plants, microorganisms, and the sediment (Kadlec and Knight, 1996). Where space allows, constructed wetlands may also be established for treating dilute aquaculture effluent. To improve effectiveness of such engineered systems, underlying factors governing nutrient removal processes need to be understood and ultimately controlled. Several studies have investigated the efficiency of subsurface- and
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Corresponding author. E-mail address: [email protected] (M. von Ahnen).
https://doi.org/10.1016/j.ecoleng.2019.105686 Received 25 June 2019; Received in revised form 8 November 2019; Accepted 20 November 2019 0925-8574/ © 2019 Elsevier B.V. All rights reserved.
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termed “wetland stream 1” and “wetland stream 2” (Fig. 1 and Table 1). Wetland stream 1 had a total FWS area of 7359 m2. It was formed as a long, narrow channel and received the production unit overflow (3474–3737 m3/d) entering at station A (Fig. 1). In addition, it received the overflow (43 m3/d, station B in Fig. 1) from small holding tanks storing drum filter reject water before pumped to the sludge tank. Wetland stream 1 was operated at a HRT of 1.5 days prior to this study (Table 2). Wetland stream 2 had a total FWS area of 11,078 m2. Water firstly entered two, interconnected basins succeeded by an alternating straight and meandering channel. Shortly before the final wetland outlet, water from wetland stream 2 mixed with water from wetland stream1 (Fig. 1, station L). Wetland stream 2 received sludge pond overflow on Sundays (314 m3; Fig. 1, station G) and biofilter backwash (144 m3/d; Fig. 1, station F) on Mondays – Saturdays (see section 2.1.1). It was operated at a hydraulic retention time of 41.0 days prior to this study calculated by averaging out discrete sludge-based input over a week.
it was argued that denitrification was limited by high oxygen concentrations and low organic carbon availability. In comparison, a two-year monitoring study in eight constructed wetlands associated with Danish MTFs type III found average removal rates of 1.5–1.9 g NO3-N/m2/d (Svendsen et al., 2008). All eight wetlands were characterized by relatively long HRTs (20–50 h), and farm effluent concentrations were higher in organic carbon and nitrate-N than in the wetland study by Dalsgaard et al. (2018) (approximately 10.0 versus 3.5 mg total BOD5/l, and 6.6 versus 3.4 mg NO3-N/l). These average removal rates are, however, still relatively low compared to other denitrification technologies applied in aquaculture (van Rijn et al., 2006) and the current, more or less passive, operation of aquaculture wetlands prevents many Danish farmers from increasing their production because of exceeding their discharge allowance (Danish Ministry of the Environment, 2012). A study by Suhr et al. (2013) demonstrated that RAS sludge may be used as a cost-free carbon source in a single-sludge denitrification process aimed at removing nitrogen from fish farm effluent. By mixing nitrate-rich effluent from a production unit with carbon-rich sludge in a heterotrophic denitrification reactor, nitrogen was removed at a maximum rate of 124.8 ± 15.7 g NO3-N/m3 reactor/d. The study highlighted that C/N ratios and HRT are essential parameters with respect to maximizing nitrogen removal in this type of setup. As an alternative (or in addition) to using a separate, rather complex filter set-up as the one by Suhr et al. (2013), we hypothesized that the sludge denitrification process may be exploited within the basins of an existing constructed wetland. To examine this, an 11-week field study was carried out with the aim of improving overall nitrogen removal in a constructed FWS wetland treating effluent from a Danish Model Trout Farm type III. The objective of the study was to mix internally generated, carbon-rich sludge effluent with increasing proportions of nitraterich effluent in a wetland side stream and measure the effect on overall nitrogen removal rates.
2.2. Experimental protocol The study was divided into two phases: 1) a “phase I" in which the original (baseline) nitrogen removal performance of the wetland was documented; and 2) a flow manipulation phase (“phase II") in which increasing proportions of nitrate-rich inflow to wetland stream 1 were re-directed to the sludge-fed wetland stream 2, and effects on wetland nitrogen removal performance evaluated. 2.2.1. Phase I–Baseline sampling Phase I documented the performance of the two wetland streams and the constructed wetland as a whole (CW) before any flow manipulations. Five, 24 h pooled sampling stations (stations A, B, C, D, and E in Fig. 1) were positioned along wetland stream 1 to follow the longitudinal turnover of nutrients and organic matter. In wetland stream 2, two 24 h pooled sampling stations (H and J in Fig. 1) were established within the wetland. In addition, to two grab sampling stations were established; one for biofilter backwash sampling (station F in Fig. 1) and one for sampling the weekly sludge pond discharge (station G in Fig. 1). Finally, a 24 h pooled sampling station was placed at the final wetland outlet (station L in Fig. 1). Baseline samples were obtained once a week for three consecutive weeks (May 10, 17, and 24, 2017; i.e., n = 3). Dissolved oxygen, pH, and temperature were measured on sample collection. Each day, 3100 kg of feed were administered to the RAS during sampling phase I.
2. Materials and methods 2.1. Study site 2.1.1. Fish farm and operation The study was carried out at a Danish MTF type III producing approximately 600 t rainbow trout/y. The inlet water to the farm consisted of ~25 l/s of groundwater and ~25 l/s of drainage water, which was recovered from below the associated constructed wetland and equally distributed into six independent RAS units. Each unit contained two parallel raceways, sludge cones, drum filters, and a fixed bed biofilter, and recirculation flows were generated via airlifts. Each biofilter was separated into six sections (24 m3 each section), each of which was backwashed on different weekdays, leaving one day per week without backwashing. Backwashing was carried out by disrupting the water flow to a section, injecting air from the bottom for about 10 min, and discharging the water from the section into an adjacent, constructed FWS wetland (Fig. 1). Sludge cones were emptied once a day, and collected sludge and drum filter backwash was pumped into a concrete sludge holding tank (r = 10 m, h = 3.5 m, V = 1100 m3). Once a week, the top 1 m dissolved layer (V = 314 m3) was discharged to a sludge pond (V = 100 m3) that fed into the adjacent wetland. The holding tank was completely emptied twice a year and the sludge transported away.
2.2.2. Phase II–Flow manipulations In study phase II, increasing proportions of nitrate-rich effluent from the production unit were re-directed at station A to the sludge-fed wetland stream 2, aiming at improving overall nitrogen removal in this wetland side stream and in the wetland as a whole. At the same time, the number of 24 h pooled sampling stations in wetland stream 2 was increased from two to five (Fig. 1 stations A, H, I, J, K) while it was reduced from five to three in wetland stream 1 (Fig. 1, stations A, B, E). Flows were manipulated on June 7, June 29 and July 13, 2017. This was done by the farm manager, increasing the overflow volume (via regulation of weir height) from station A into wetland stream 2 from 0 to 1431, 1773, and 2015 m3/d, respectively. The redirected flows corresponded to approximately 40, 49, and 54% of the former wetland stream 1 inflow, and are henceforth referred to as flow 1, 2, and 3. The aim was to establish three increasing flows up to the maximum possible gravity flow in wetland stream 2, without manipulating the weir that delivered water to wetland stream 1. Flow manipulations were succeeded by acclimatization periods of 20, 12, and 12 days, respectively, prior to sampling, corresponding to 5.3, 4.0, and 4.5 times the retention time in wetland stream 2. Pooled 24 h samples and grab samples were obtained for three consecutive days (i.e., n = 3) following each acclimatization period. The farm fed
2.1.2. Constructed wetland A constructed wetland with a total FWS area of 19,007 m2 and an average depth of 0.7 m was associated with the farm for effluent polishing before final discharge to a local stream. The wetland had been in use for 10 years prior to the study, and was constructed by interconnecting former earthen fish ponds (emptied for fish and no longer in used for fish farming) into two, parallel wetland streams henceforth 2
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Figure 1. Overview of the constructed free water surface wetland treating the effluent from a recirculating Model Trout Farm type III. Indicated is the monitored stretches of the two wetland streams referred to as wetland stream 1 (blue) and wetland stream 2 (red), direction of water flow, and sampling stations (A-L). Orange lines indicate non-monitored stretches prior to the shared outlet (final discharge from the entire constructed wetland) at monitoring station L. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
using automatic, refrigerated samplers (Glacier® Portable, Teledyne ISCO, USA) programmed to sample 300 ml with 1 h intervals for 24 h. Subsamples were filtered through 0.45 μm mixed cellulose ester filters (Whatman, GE Healthcare, UK) prior to analyses of: dissolved five-day biochemical oxygen demand (BOD5-diss; ISO 5815-2, 2003); dissolved chemical oxygen demand (CODdiss.; ISO 15705, 2002; 4–40 mg/l, 15–150 mg/l, 150–1000 mg/l Spectroquant® test kits, Merck Millipore, Germany); and orthophosphate (PO4-P). Other subsamples were filtered through 0.2 μm syringe filters (Filtropur S 0.2 μm, Sarstedt, Germany) prior to analyses of: total ammonia nitrogen (TAN; DS 224, 1975); nitrite-N (NO2-N; DS 223, 1991); nitrate-N (NO3-N; ISO 7890–1, 1986); and dissolved total N (TNdiss; ISO 7890–1, 1986; ISO 11905–1, 1997). Finally, unfiltered sub-samples were analyzed for total BOD5 (BOD5-total; ISO 5815-2, 2003); total COD (CODtotal; ISO 15705, 2002; 4–40 mg/l, 15–150 mg/l, 150–1000 mg/l Spectroquant® test kits, Merck Millipore, Germany); total N (TN; ISO 7890–1, 1986; ISO 119051, 1997); and total phosphorous (TP; ISO 6878, 2004). Samples were kept cold (3 °C) until analysis. Analyses of BOD5 were initiated on the day of sample collection while all other analyses were performed within the first few days after collection. Unfiltered and filtered subsamples for CODtotal, CODdiss, PO4-P, and TP analyses were conserved until analysis by adding 1% sulfuric acid (H2SO4; 21.4%; Merck KGaA, Germany). Dissolved oxygen (mg O2/l), pH, and temperature (°C) were measured at each sampling station using Hach Lange HQ40 multimeters (Hach Lange, Germany).
Table 1 Wetland sampling stations cf. Fig. 1 positioned at increasing, free water surface (FWS) distance from the inlet. Wetland section
Sampling station
FWS area (m2)
Wetland stream 1
A (production unit effluent) B (drum filter reject water) C D E F (biofilter backwash) G (sludge overflow) H I J K L (wetland outlet)
0 0 2934 4625 7359 0 0 2512 5138 6293 8344 19,007
Wetland stream 2
Entire wetland (CW)
2800, 2600, and 3100 kg feed per day, respectively, during the three sampling periods. Flow rates into wetland stream 1 and 2 were derived each sampling day by measuring weir overflow heights and applying Torricelli's law (Molitor, 1908). Additional inflow contributions at station B were measured at an open pipe using a bucket and a stop watch, while inflows at stations F and G were calculated (geometrically) from known tank volumes discharged into the wetland. Wetland outflow rates were assumed to equal inflow rates, i.e., infiltration, precipitation, evapotranspiration, etc. in the wetland was assumed to be negligible. 2.3. Sample analyses
2.4. Calculation of removal rates and statistical analyses Area-based removal rates (g/m2/d) were calculated as the mass
Pooled samples were collected from the centre of the water column
Table 2 Flow rates, hydraulic retention times (HRTs), and influent C/N ratios (as CODdiss/NO3-N) measured during sampling phase I and II. Values in same columns not sharing a common superscript are significantly different (P < .05). Wetland stream 1
Wetland stream 2
Sampling phase
Inflow st. A (m3/ d)
Inflow st. B (m3/ d)
HRT (d)
Influent C/N ratio
Inflow st. A (m3/ d)
Inflow st. F⁎ (m3/ d)
Inflow st. G⁎ (m3/ d)
HRT (d)
Influent C/N ratio
Phase Phase Phase Phase
3474 2146 1860 1722
43 43 43 43
1.5 2.4 2.7 2.9
1.4 1.3 1.1 1.1
0 1431 1773 2015
144 144 144 144
45 45 45 45
23.3 2.7 2.3 2.0
1086 ± 57a 216 ± 54b 233 ± 55b 261 ± 49b
⁎
I, baseline II, flow 1 II, flow 2 II, flow 3
± ± ± ±
0.3 0.1 0.2 0.2
Calculated by averaging out discrete sludge-based input over a week. 3
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Table 3 Removal rates in g/m2/d⁎ (mean ± SD, n = 3, for BOD5n = 2) of various nutrients for wetland stream 1 (W1), wetland stream 2 (W2)⁎⁎, and the entire constructed wetland (CW) as well as average wetland temperatures (°C) obtained during measurements of the baseline performance (phase I) and the flow manipulations (flows 1–3, phase II). Nutrient
Wetland
Baseline
Flow 1
Flow 2
Flow 3
TN
W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW W1 W2 CW CW
1.4 ± 0.4 0.1 ± 0.0a 1.4 ± 0.2a 1.2 ± 0.6 1.7 ± 0.1a 1.4 ± 0.2a 1.1 ± 0.2 0.3 ± 0.1a 0.3 ± 0.6a 0.13 ± 0.05 0.02 ± 0.01 0.08 ± 0.03 0.2 ± 0.2 1.6 ± 0.2a 0.5 ± 0.2 0.02 ± 0.01ab 1.00 ± 0.02a 0.27 ± 0.08a 0.03 ± 0.03 0.62 ± 0.05a 0.15 ± 0.06a 0.4 ± 2.6 45.1 ± 1.4 16.0 ± 1.0 2.0 ± 0.5 30.5 ± 0.9a 10.7 ± 0.2a 1.0 ± 1.2 21.7 ± 14.1 7.2 ± 4.2 1.4 ± 0.3 14.1 ± 9.3 5.3 ± 3.2a 13.3 ± 1.8 a
1.1 ± 0.4 4.8 ± 0.3b 2.2 ± 0.2b 1.1 ± 0.3 2.3 ± 0.2ab 1.3 ± 0.2ab 0.8 ± 0.3 3.1 ± 0.4b 1.4 ± 0.1b 0.05 ± 0.04 −0.12 ± 0.13 0.01 ± 0.01 0.1 ± 0.1 1.8 ± 0.1a 0.2 ± 0.6 0.01 ± 0.02a 1.42 ± 0.03b 0.54 ± 0.03b −0.01 ± 0.01 1.27 ± 0.11b 0.46 ± 0.06b 2.3 ± 0.6 91.7 ± 34.5 32.4 ± 12.3 1.7 ± 0.3 59.3 ± 1.4ab 20.6 ± 0.9b 2.2 ± 0.4 35.1 ± 4.9 16.7 ± 2.4 1.3 ± 0.1 34.4 ± 1.7 15.8 ± 0.9b 14.8 ± 0.8 b
1.4 ± 0.1 5.5 ± 0.1c 2.8 ± 0.1c 1.4 ± 0.2 2.2 ± 0.1ab 1.7 ± 0.1ab 1.0 ± 0.1 3.6 ± 0.3b 2.1 ± 0.1b 0.01 ± 0.02 −0.08 ± 0.02 0.02 ± 0.01 0.1 ± 0.0 1.9 ± 0.2a 0.6 ± 0.0 0.04 ± 0.01ab 1.49 ± 0.11c 0.43 ± 0.02ab 0.01 ± 0.01 1.27 ± 0.12c 0.38 ± 0.04ab 1.8 ± 0.6 73.0 ± 0.4 25.0 ± 0.4 1.4 ± 0.5 64.2 ± 1.3ab 21.9 ± 0.6c 1.3 ± 0.2 40.5 ± 1.8 19.4 ± 0.2 0.7 ± 0.3 39.2 ± 0.6 17.9 ± 0.8b 15.4 ± 0.7 c
1.8 ± 1.0 8.4 ± 1.4d 3.9 ± 0.8c 1.7 ± 1.1 7.4 ± 1.1b 3.7 ± 0.9b 1.2 ± 1.1 5.1 ± 1.7b 2.7 ± 1.1b 0.03 ± 0.04 0.00 ± 0.05 0.04 ± 0.02 0.2 ± 0.2 3.4 ± 0.3b 1.1 ± 0.3 0.06 ± 0.01b 1.99 ± 0.05d 0.56 ± 0.08b 0.02 ± 0.02 1.88 ± 0.33d 0.54 ± 0.16b 2.3 ± 0.5 98.9 ± 1.1 33.2 ± 0.3 1.5 ± 0.3 86.8 ± 6.8b 29.3 ± 2.4d 1.7 ± 0.4 43.9 ± 1.6 20.4 ± 0.5 0.9 ± 0.4 43.6 ± 0.9 19.8 ± 0.7b 16.0 ± 0.7 c
TNdiss
NO3-N
NO2-N
TAN
TP
PO4-P
CODtotal
CODdiss
BOD5-total
BOD5-diss
Temperature
⁎ Values in the same row not sharing a common superscript letter are significantly different (P < .05). The absence of letters in rows indicates that values are not significantly different. ⁎⁎ Note that removal rates for wetland 2 are based on the area between the wetland inlet and station J as station K was not sampled during baseline sampling (section 2.1.2).
comparison, there was no change in total nitrogen removal in wetland stream 1, ranging between 1.1 ± 0.4 and 1.8 ± 1.0 g TN/m2/d throughout the study (Table 3). These results sustain that denitrification, prior to flow manipulations, was nitrate-limited in wetland stream 2, while it was limited by low carbon availability and high oxygen concentrations in wetland stream 1 (Fig. 2e) as previously observed in an aquaculture wetland operated at short HRT (Dalsgaard et al., 2018). Hence, nitrate limitation typically occurs at NO3-N concentrations below 1 mg/l (Henze et al., 2002). Consistent with this, an average NO3-N concentration of 0.6 ± 0.3 mg/l was measured at station J in wetland stream 2 during baseline sampling. At such low NO3-N concentrations, available organic matter typically fuels heterotrophic bacterial processes such as sulfate reduction and methanogenesis rather than heterotrophic denitrification (Korom, 1992). During flow manipulations, the nitrate load on wetland stream 2 increased from a baseline value of 0.3 ± 0.2 to 4.6 ± 0.3, 5.8 ± 0.1, and 7.4 ± 1.6 g NO3-N/m2/d, respectively, at flow 1, 2 and 3. Corresponding NO3-N concentrations at station J increased from a baseline value of 0.6 ± 0.3 to 5.6 ± 0.4, 6.8 ± 0.8, and 6.7 ± 0.5 mg/l. I.e., the flow manipulations apparently improved the conditions for heterotrophic denitrification in wetland stream 2 to an extent that it was no longer nitrate limited. According to Fig. 2a, most of the nitrogen removal took place prior to station H, inferred from the decline in TN removal rates following this station. Organic matter removal followed a similar pattern (Fig. 2 c,
removed between the wetland inlet station and a specific wetland station (assuming plug flow conditions) and dividing with the FWS wetland area between the two stations. Results are throughout shown as mean ± SD. Statistical analyses were carried out in SigmaPlot version 13.0 (Systat Software Inc., CA, USA) applying a significance level of p < .05. Removal rates following flow manipulations were compared using One-way ANOVA followed by Holm Šidák multiple comparison test in case of significant differences. In case the normality test (Shapiro-Wilk) failed, one-way ANOVA on ranks was run followed by a Tukey multiple comparison test. Spearman correlation analysis was applied for testing the correlation between area-based NO3-N and CODdiss removal rates.
3. Results and discussion 3.1. Nitrogen removal Increasing the inflow of nitrate-rich effluent to wetland stream 2 led to a significant increase in overall nitrogen removal of the entire wetland (CW; Tables 3 and 4). The area-based TN removal rate for the whole wetland increased from a baseline value of 1.4 ± 0.2 to 3.9 ± 0.8 g/m2/d at the highest manipulated flow (flow 3). This improvement was principally due to a significant increase in wetland stream 2 nitrate removal (in area delimited by station J), increasing from a baseline value of 0.3 ± 0.1 to 5.1 ± 1.7 g NO3-N/m2/d. In 4
Ecological Engineering 143 (2020) 105686
55 55 60 31 55 60 65 86 89 93 97 15.5 ± 0.4 14.4 ± 0.9 8.7 ± 1.0 0.43 ± 0.01 4.37 ± 0.55 1.80 ± 0.41 1.37 ± 0.28 26.0 ± 3.5 18.1 ± 0.3 7.0 ± 2.7 2.6 ± 0.2
d), and there was a positive correlation (r = 0.66, p < .001, n = 36) between NO3-N and CODdiss removal rates in the flow manipulation phase. It thus seems that heterotrophic denitrification switched from being nitrate limited prior to flow manipulations to possibly becoming carbon limited downstream wetland stream 2. 3.2. Organic matter removal Organic matter removal rates of both COD and BOD5 were at all times considerably higher in wetland stream 2 than in wetland stream 1 (Table 3), reflecting different hydraulic conditions and organic matter loadings. Flow manipulations led to a significant increase in CODdiss removal rates in wetland stream 2, increasing from a baseline value of 30.5 to 86.8 g/m2/d at the highest manipulated flow. The removal rate of BOD5-diss similarly increased (from 14.1 to 43.6 g/m2/d; Table 3), however, the increase was not statistically significant presumably due to the low number of replicates (n = 2). These improvements in dissolved organic matter removal rates are consistent with the higher denitrification activity observed in wetland stream 2 as discussed in section 3.1. In addition to carbon from added sludge overflow, additional carbon may derive from wetland-internally generated, dead and decomposing plant biomass. However, not all of this carbon is readily available for denitrifiers, and previous wetland studies have shown that carbon can be limiting at high nitrate loading, in which case denitrification will depend strongly on the carbon supply rate and availability (reviewed by Kadlec and Wallace, 2009; Hang et al., 2016). There were no significant changes in organic matter removal rates in wetland stream 1 following flow manipulations. However, total organic matter removal rates were in all cases slightly higher and dissolved organic removal rates slightly lower than at baseline (Table 3). Flow manipulations increased the hydraulic retention time in wetland stream 1 from 1.5 to 2.9 days (Table 2). The time available for breakdown of particulate biodegradable organic matter thereby increased, and this may explain the tendency for a higher area-based removal rate. Dissolved organic matter, on the other hand, was probably fully degraded in wetland stream 1 even before flow manipulations given the high oxygen concentrations (Fig. 2e), and a longer retention time therefore actually reduced the area-based removal rate of dissolved organic matter in wetland stream 1. 3.3. C/N ratios
38 28 37 8 15 72 72 88 86 95 99 28.6 ± 2.3 23.0 ± 1.4 18.6 ± 1.3 0.70 ± 0.08 6.78 ± 0.19 3.75 ± 0.10 3.18 ± 0.24 184.0 ± 60.9 120.0 ± 4.8 87.3 ± 12.1 80.2 ± 4.9
17.6 ± 1.9 16.6 ± 1.0 11.6 ± 1.5 0.64 ± 0.01 5.76 ± 2.72 1.07 ± 0.06 0.88 ± 0.04 22.2 ± 1.5 17.4 ± 1.1 4.0 ± 0.1 1.2 ± 0.3
29.0 ± 0.5 23.0 ± 0.7 19.1 ± 0.3 0.44 ± 0.01 6.91 ± 0.10 3.84 ± 0.08 3.14 ± 0.13 148.1 ± 3.0 125.0 ± 3.7 100.3 ± 2.0 90.4 ± 4.0
15.2 ± 0.1 14.4 ± 0.2 8.7 ± 0.4 0.35 ± 0.04 3.76 ± 0.06 1.71 ± 0.16 1.28 ± 0.07 26.1 ± 3.4 17.3 ± 0.9 4.8 ± 1.0 2.3 ± 0.3
48 37 54 22 46 55 59 82 86 95 97
34.3 ± 3.5 32.0 ± 3.6 21.7 ± 4.6 0.62 ± 0.12 9.75 ± 0.77 4.47 ± 0.09 3.97 ± 0.48 185.2 ± 2.2 158.6 ± 12.0 104.4 ± 0.1 97.5 ± 3.4
Removal% Outlet mg/l Inlet mg/l Removal % Outlet mg/l Inlet mg/l Removal % Outlet mg/l
27 28 12 46 28 50 42 80 77 88 95
Increasing the nitrate load on wetland stream 2 reduced the incoming C/N ratio (calculated as CODdiss/NO3-N), declining significantly from an average of 1086 ± 57 before flow manipulations to an average of 234 ± 56 following flow manipulations (Table 2). Wetland stream 2 presumably received large amounts of easily biodegradable organic matter including volatile fatty acids (VFAs) deriving from hydrolysis in the sludge tank. Prior to flow manipulations, this organic matter was likely “wasted” to aerobic respiration and subsequently to anaerobic respiration such as sulfate reduction and methanogenesis in the lack of nitrate. Introducing nitrate-rich effluent improved the conditions for denitrification, and denitrification activity was high as long as easily decomposable organic matter was available. Moving downstream in wetland stream 2, the fraction of less biodegradable organic carbon, however, increased as also observed by Puigagut et al. (2008) in a study on organic matter in a horizontal subsurface-flow wetland. Relatively high NO3-N concentrations of 6.35 ± 0.52 mg/l were thus measured at station J by the end of the study despite otherwise favorable denitrification conditions including oxygen concentrations below 0.4 mg/l (Fig. 2e) at a C/N ratio of 3.01.
26.0 ± 0.9 25.0 ± 1.5 13.9 ± 1.7 0.85 ± 0.11 8.28 ± 0.50 2.70 ± 0.02 1.85 ± 0.10 101.1 ± 2.8 71.0 ± 0.3 61.6 ± 2.7 28.7 ± 16.4 TN TNdiss NO3-N NO2-N TAN TP PO4-P CODtotal CODdiss BOD5-total BOD5-diss
19.0 ± 1.7 18.0 ± 1.9 12.2 ± 1.4 0.46 ± 0.04 5.99 ± 0.83 1.34 ± 0.40 1.07 ± 0.32 20.6 ± 2.9 16.3 ± 1.4 7.19 ± 4.3 1.4 ± 0.1
Inlet mg/l Parameter
Oultet mg/l
Removal %
Inlet mg/l
Flow 2 Flow 1 Baseline
Table 4 Averaged inlet and outlet concentrations as well as percentage removals during baseline sampling and flow manipulations for the entire constructed wetland (CW).
Flow 3
M. von Ahnen, et al.
3.4. Comparison with previous wetland denitrification studies The highest TN removal rates in the current study, 3.3 g/m2/d for 5
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a)
b) Stations:
H
I
J
K
Stations
H
I
J
Baseline Flow 1 Flow 2 Flow 3
10
a a
b
b
b b
5
b
b b
b b
b
a
Baseline Flow 1 Flow 2 Flow 3
2
a
15
0
15 a
10 a a
a b ab
5
a -5 2000
4000
6000
8000
10000
0
2000
a
b ab
b b
a 4000
Wetland stream 2 area (m2)
6000
8000
10
2
Wetland stream 2 area (m )
c)
d) H
I
J Baseline Flow 1 Flow 2 Flow 3
300 2
250
350
K
a a a
150 a
100
ab a c
b
50
b a d
b
H
I
J
Baseline Flow 1 Flow 2 Flow 3
250 a a
200
b
150
bc b a
a 100
c
c a
d a
b
50
0
K
a
300
200
ab a
Stations
2
Stations
COD diss removal rate [g/m /d]
350
CODtotal removal rate [g/m /d]
a
0 0
0 0
2000
4000
6000
8000
10
0
Wetland stream 2 area (m2)
e)
K
20
NO3-N removal rate [g/m /d]
2
TNunf il. removal rate [g/m /d]
20
2000
4000
6000
8000
10
Wetland stream 2 area (m2)
f)
Fig. 2. Area-based removal rates in g/m2/d (mean ± SD) for: a) TN; b) NO3-N; c) CODtotal; and d) CODdiss obtained during baseline sampling and following the three flow manipulations periods in wetland stream 2. Also shown is: e) dissolved oxygen concentrations; and f) pH along wetland stream 1 (W1), stream 2 (W2), and the entire wetland (CW). Area-based removal rates are expressed continuously, i.e. removal rates were calculated for the area between the wetland inlet and the respective sampling station with station letters referring to Fig. 1. Values sharing identical symbols but different lower case letters are significantly different (P < .05). Influent oxygen concentrations to wetland stream 1 and 2 are based on weighted averages.
6
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the wetland as a whole and 8.4 g TN/m2/d for wetland stream 2, may be compared with the 1.5–1.9 g TN/m2/d measured by Svendsen et al. (2008) in a monitoring study of eight aquaculture wetlands. The eight wetlands were all characterized by HRTs of 20–50 h and average inlet COD/(NO2-N + NO3-N) ratios of 5.5. All discharge from the production units was typically mixed with sludge basin overflow prior to the wetlands, explaining the lower C/N ratio compared to the current study. These different results demonstrate that fish sludge, with little effort, may be used as an internal carbon source to improve denitrification in aquaculture wetlands given the right conditions. This includes optimizing the C/N ratio (5 or 10 to 1; reviewed by Kadlec and Wallace, 2009) using recently hydrolyzed sludge, establishing anaerobic zones, and treating only a side stream of the production unit discharge to sustain that easily biodegradable organic matter is not “wasted” in aerobic respiration. The latter was probably the situation in a FWS wetland treating dilute effluent from a Danish recirculating farm at short retention time (0.32 d; Dalsgaard et al., 2018). No nitrate removal was observed in that study during a whole year of monitoring. Rather, the wetland even became a net nitrate producer at certain times presumably due to a lack of readily available organic carbon combined with high oxygen concentrations. Similarly, Schulz et al. (2004) found comparatively low NO3-N removal rates of 0.06 ± 0.14 g/m2/d at a HRT of 11 h, and a net production of up to 0.27 ± 0.12 g NO3-N/m2/d at a HRT of 3.5 h in FWS wetlands treating highly dilute effluent from a flow-through rainbow trout fingerling farm. On the other hand, when too little production effluent mixes with available sludge effluent in a wetland, nitrate limitation may reduce overall nitrogen removal as was the case in wetland stream 2 during baseline sampling (very high C/N ratio).
majority of excreted phosphorous will be bound in solid waste (Dalsgaard and Pedersen, 2011). Solid waste can be removed by sedimentation (Kadlec and Wallace, 2009), and sedimentation of suspended particles may explain some of the higher removal in wetland stream 2. Sedimentation of particulate P probably also explains the similar removal rates obtained by Schulz et al. (2004), studying removal rates in three parallel FWS aquaculture wetlands treating effluent from a flowthrough farm. In that study, TP removal rates increased from 0.12 to 0.30 g TP/m2/d when loadings increased from 0.23 to 0.70 g TP/m2/d. In addition to particle sedimentation, higher denitrification activity in wetland stream 2 may also have improved phosphorous removal. Certain denitrifiers are known to accumulate orthophosphate in excess of their metabolic requirements leading to a considerable reduction in phosphorous load in the water column (van Rijn et al., 2006). In addition, heterotrophic denitrification activity produces alkalinity (van Rijn et al., 2006). Combined with a higher effluent volume entering wetland stream 2, this presumably contributed to raising outlet pH in wetland stream 2 from a baseline value of 6.8 to above 7.0 (p < .05; Fig. 2f). An increase in pH may cause calcium phosphate to precipitate in denitrifying biofilms in which pH can locally increase above values measured in the water (Arvin and Kristensen, 1982). In comparison to wetland stream 2, wetland stream 1 received low concentrations of primarily dissolved phosphorous. Dissolved phosphorous may be removed by sorption, plant and microbial uptake, and sediment accretion (Kadlec and Wallace, 2009). However, at very low concentrations (approaching a non-zero background concentration) there will be no net removal due to internal cycling, including return flux from decomposition processes, and atmospheric gains and losses (Kadlec and Wallace, 2009). This situation was probably the case in wetland stream 1 where almost no P-removal was observed.
3.5. TAN removal
4. Summary and future prospects
Flow manipulations led to a significant increase in the area-based TAN removal rate in wetland stream 2, increasing from a baseline value of 1.6 to 3.4 g/m2/d at the highest flow (Table 3). No similar effect was observed in wetland stream 1, while for the wetland as a whole there was a non-significant increase from a baseline value of 0.5 to 1.1 g/m2/ d (Table 3). The increase in TAN removal in wetland stream 2 may relate to the reduction in the incoming C/N ratio (Table 2), possibly leading to a reduction in DNRA activity (dissimilatory nitrate reduction to ammonia). This process is conducted by fermentative bacteria in excessive reductive environments (i.e., high C/N; Washbourne et al., 2011) as prevailing in wetland stream 2 especially during baseline sampling. A reduction in the C/N ratio may also have favored an increase in autotrophic anaerobic ammonia oxidation (anammox activity). This anaerobic and autotrophic microbial process converts ammonia and nitrite to dinitrogen, and may have lead to additional removal of TAN and TN (Mulder et al., 1995; Zhu et al., 2011; Suhr et al., 2013).
This case study demonstrated that nitrogen removal in an aquaculture wetland can be improved by using internally generated sludge/ fish waste as a carbon source for denitrification in an anoxic part of the wetland. As a side effect, some additional removal of phosphorous may be achieved via the enhanced denitrification activity. Through flow manipulations, it was demonstrated that the farm was able to significantly reduce the discharge of nitrogen and its overall environmental impact(Table 4). For further optimization, controlled biological sludge hydrolysis prior to the wetland, i.e., solubilization and fermentation of settled organic matter to volatile fatty acids (VFAs), may improve the denitrification process (Letelier-Gordo et al., 2015; Letelier-Gordo et al., 2017). Sludge hydrolysis is carried out by naturally occurring facultative anaerobic bacteria, and happens by itself if particulate organic matter is left to settle. However, solubilized organic matter should be transferred to the wetland after a few days as the anaerobic degradation process otherwise continues into a methanogenic stage and degradable carbon will be “lost” as methane (Letelier-Gordo et al., 2015, LetelierGordo et al., 2017). In this particular case study, wetland stream 1 carried a side stream high in nitrate and low in organic carbon. This combination is ideal for further denitrification treatment in a woodchip bioreactor recently demonstrated as another low-cost, end-of-pipe denitrification technology successfully applied in aquaculture (von Ahnen et al., 2018).
3.6. Phosphorous removal Besides improving nitrogen and organic matter removal, flow manipulations also improved the removal rates of TP and PO4-P in wetland stream 2 increasing significantly from 1.00 ± 0.02 to 1.99 ± 0.05, and from 0.62 ± 0.05 to 1.88 ± 0.33 g/m2/d, respectively (Table 3). No changes were observed in wetland stream 1 apart from an infinitesimal improvement in TP removal. Considering the wetland as a whole, there was a relatively small but significant improvement in TP and PO4-P removal (Table 4). Wetland stream 2 received a higher total phosphorous load than wetland stream 1 (1.13 ± 0.02 vs. 0.39 ± 0.03 g TP/m2/d, baseline sampling) including a relatively larger share of particle bound phosphorous from the sludge/particulate waste. When feeding a commercial diet balanced in phosphorous (as in the current case study), the
Authors contributions Mathis von Ahnen participated in the design of the study, carried out the sampling and drafted the manuscript. Per Bovbjerg Pedersen participated in the design of the study and helped to draft the manuscript. Johanne Dalsgaard participated in the design of the study and coordination and helped to draft the manuscript. All authors read and approved the final manuscript. 7
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Declaration of Competing Interest
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The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgements This research was funded by the Ministry of Food, Agriculture and Fisheries of Denmark and by the European Union through the European Maritime and Fisheries Fund (EMFF). We thank the fish farmer for allowing us access to his facilities and the farm manager for helpful assistance during the study. In addition, we thank laboratory technician Ulla Sproegel, Brian Møller, and Dorthe Frandsen (DTU Aqua) for invaluable technical assistance in the laboratories. References Arvin, E., Kristensen, G.H., 1982. Precipitation of calcium phosphate and pH-effects in denitrifying biofilms. Water Sci. Technol. 14 (9–11), 1505–1508. Dalsgaard, J., Pedersen, P.B., 2011. Solid and suspended/dissolved waste (N,P,O) from rainbow trout (Oncorynchus mykiss). Aquaculture 313, 92–99. Dalsgaard, J., von Ahnen, M., Naas, C., Pedersen, P.B., 2018. Nutrient removal in a constructed wetland treating aquaculture effluent at short hydraulic retention time. Aquaculture Environment Interactions 10, 329–343. Danish Ministry of the Environment, 2012. Bekendtgørelse om miljøgodkendelse og samtidig sagsbehandling af ferskvandsdambrug (“in Danish”; Executive order on the environmental approval and handling of freshwater aquaculture systems). Lovtidende A nr. 130. Miljøstyrelsen, Miljøministeriet, Denmark. DS 223, 1991. Water Analysis—Determination of the Sum of Nitrite- and NitrateNitrogen. Danish Standards Foundation, Charlottenlund, Denmark. DS 224, 1975. Water Analysis—Determination of Ammonia-Nitrogen. Danish Standards Foundation, Charlottenlund, Denmark. Hang, Q., Wang, H., Chu, Z., Ye, B., Li, C., Hou, Z., 2016. Application of plant carbon source for denitrification by constructed wetland and bioreactor: review of recent development. Environ. Sci. Pollut. Res. Int. 23, 8260–8274. Henze, M., Harremoës, P., Jes la Coeur, J., Arvin, E., 2002. Wastewater Treatment: Biological and Chemical Processes, 3rd edition. Springer Verlag, Berlin Heidelberg, Germany. ISO 11905-1, 1997. Water Quality—Determination of nitrogen. Part 1: Method Using Oxidative Digestion with Peroxide Sulfate. International Organization for Standardization, Geneva, Switzerland. ISO 15705, 2002. Water Quality—Determination of the Chemical Oxygen Demand Index (ST-COD)—Small-Scale Sealed-Tube Method. International Organization for Standardization, Geneva, Switzerland. ISO 5815-2, 2003. Water Quality—Determination of Biochemical Oxygen Demand after N Days (BODn)—Part 2: Method for Undiluted Samples, ISO 5815-2:2003, Modified. International Organization for Standardization, Geneva, Switzerland. ISO 6878, 2004. Water Quality—Determination of Phosphorus—Ammonium Molybdate Spectrometric Method. International Organization for Standardization, Geneva, Switzerland.
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