Application of a Full-scale Constructed Wetland for Tertiary Treatment ...

Water Air Soil Pollut (2008) 193:15–24 DOI 10.1007/s11270-008-9664-5

Application of a Full-scale Constructed Wetland for Tertiary Treatment of Piggery Manure: Monitoring Results E. Meers & F. M. G. Tack & I. Tolpe & E. Michels

Received: 30 September 2007 / Accepted: 1 March 2008 / Published online: 3 April 2008 # Springer Science + Business Media B.V. 2008

Abstract Many industrialized regions in the world are faced with local overproductions of animal manure requiring processing in an economic sound manner. Intensive animal production in Flanders and the Netherlands has resulted in a considerable overproduction of animal manure. Spreading the excess manure over arable land has resulted in contamination and eutrophication of groundwater and surface waters. Over the last 4 years, research was conducted towards the potential of more economic constructed wetlands for the final treatment step. Although, initial results with laboratory flow field experiments were insufficient to reach stringent discharge criteria (Meers et al., Water Air Soil Pollut 160:15–26, 2005a), progressive optimisation of the tertiary treatment as well as of the preceding conditioning has resulted in a consistently performing pilot scale system (1,000 m3 year−1 capacity) with effluent concentrations below the discharge criteria of 15 mg l−1 N, 2 mg l−1 P and 125 mg l−1 COD (chemical oxygen demand), at a cumulated cost (operational plus investment) of 3–4 € m−3 of pre-treated pig manure. Construction of full-scale installations with annual capacity of 10,000–25,000 m3 based on this pilot E. Meers (*) : F. M. G. Tack : I. Tolpe : E. Michels Department of Applied Analytical and Physical Chemistry, Laboratory of Analytical Chemistry and Applied Ecochemistry, Ghent University, Coupure Links 653, 9000 Gent, Belgium e-mail: [email protected]

model are scheduled, with the first installation currently under way. The concept has the potential to provide a low cost, in situ treatment system allowing animal farmers to process excess animal manure themselves without the requirement of expensive ex situ treatment based on industrial scale membrane technology facilities. This paper presents the research findings of the first year of the pilot scale installation. Keywords Constructed wetlands . Manure . Manure treatment . Manure processing . Wastewater treatment . N

1 Introduction In many industrialized areas of the world, manure production originating from intensive animal farming exceeds the regional capacity for use on arable land as a fertilizer. Flanders (Belgium) has a production of animal manure containing 158 million kg of N while only 96 million kg are needed for fertilization. This imbalance in production versus demand has contributed to strongly increased nitrate concentrations in both groundwater and surface water. In 2005–2006, 42% of selected sampling locations located in agricultural areas exhibited water concentrations exceeding 50 mg l−1 NO3, which is the maximum allowable limit according to the European nitrates directive (91/676/EEC). Water-soluble N and P compounds are known to stimulate eutrophication of

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surface waters, resulting in extreme fluctuations of the dissolved oxygen concentrations with detrimental effects on aquatic organisms. Enhanced nitrate concentrations in groundwater can cause a significantly increased production cost of potable water. Emission of nitrate into the environment may therefore result in considerable ecological, social as well as economical effects (Hill and Sobsey 2001; Hunt and Poach 2001). In September 2005, Belgium was condemned by the European Union for not respecting the nitrate directive and imposed a number of sanctions and demands for improvement of water quality. In 2007, a new national legal framework was adopted in which pig manure processing was stimulated. The current state of the art in manure processing in Flanders (Belgium) involves three phases: (a) separation into liquid and solid fractions, (b) conversion of the solid fraction into an exportable product via e.g. composting and (c) reduction of the nutrient content in the liquid fraction either to below discharge criteria or to levels suitable for application on arable land. Currently, in Flanders and other European regions with intensive animal farming the most widely used technology for N reduction in the liquid fraction involves secondary treatment in a biological treatment reactor (Smet et al. 2003). This technology is currently also being introduced in the United States, as can be illustrated by the pilot-scale installation in North-Carolina described by Vanotti et al. (2007), which has a close resemblance to the European manure treatment facilities. Meers et al. (2005a) proposed the use of horizontal subsurface flow beds (SSF) with Phragmites australis to reduce nutrient concentrations to below discharge criteria (15 mg/l N, 2 mg/l P), after secondary treatment in an activated sludge reactor. Hunt and Poach (2001) and Poach et al. (2003) also recommended having pre-wetland procedures that promote oxidation when high mass removals of N and P are required. Knight et al. (2000) stated that wetland design for treatment of livestock wastewater, including piggery manure, must include adequate pretreatment to protect the health of the wetland biota and to meet the quality goals. In the wetland experiments described by Meers et al. (2005a), effluent levels of pre-treated piggery manure still remained significantly above the Flemish legal discharge criteria of 2 mg l−1 P, 15 mg l−1 N and 125 mg l−1 COD. This was much in accordance with

Water Air Soil Pollut (2008) 193:15–24

Knight et al. (2000) who observed average outflow concentrations of as high as 248 mg l−1 N for treatment wetlands designed for comparable livestock wastewater (including piggery manure) based on an elaborate comparison of 135 pilot and full-scale wetland systems. Poach et al. (2003) also described improved N removal in liquid swine manure by wetlands after partial nitrification yet also not to the point where effluent N concentrations below 15 mg l−1 were achieved. Since then, the experiments by Meers and co-workers have been scaled up and moved to outdoor conditions. The wetland design has been upgraded to a multi-bed system, including a cascade of horizontal and vertical helophyte filters as well as hydrophyte and pleustophyte lagoons. The upgraded system has been transformed into a pilot-scale installation designed for processing 1,000 metric tons of pig manure per year. Construction of this installation was performed in spring 2006. In autumn, after 4 months of operation the system proved to be sufficiently robust and the wetlands were further expanded to allow a capacity of 5,000 m3 year−1 from the growing season of 2007 onwards. Finally, in spring 2007 construction of the first full-scale installation, capable of processing 25,000 m3 year−1 was initiated. Seven further projects of this scale scheduled for construction in 2007 and 2008. This paper wishes to extend the monitoring results of the pilot-scale installation to the scientific community to illustrate that using a chain of environmental technologies, with constructed wetlands as a concluding step, allows to transform the liquid fraction of pig manure into re-usable/dischargeable water. To our knowledge, this is the first successful application of this technology to convert manure into water of sufficient quality regarding N, P and COD content at this scale.

2 Material and Methods 2.1 Wetland Overview The system involves a multi-bed system, existing out of eight beds/lagoons, with both vertical and horizontal flow helophyte beds, pleustophyte ponds and hydrophyte lagoons. These include two horizontal flow fields, two vertical percolation fields, two underwater systems and two lagoons with floating


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plant systems. A schematic overview of the constructed wetland is depicted in Fig. 1. The total surface area of the pilot-scale system is 4,500 m2. The system is divided into two parts, an intensive zone (1,000 m2) and an extensive zone (3,500 m2). In the intensive zone, nutrient and COD concentrations are high, pump regimes are intensive and substrates and materials are more expensive yet also more performing. The extensive zone, designed for the final polishing, is a plug flow system with gravitational flow, lower nutrient and COD concentrations and more economic substrates, however spread over a larger surface area. The helophyte filter beds contain a diverse mixture of among others P. australis, Typha latifolia, Carex pseudocyperus, Carex acutiformis, Scirpus lacustris, Scirpus maritimus, Filipendula ulmaria, Iris pseudacorus, Mentha aquatica, Acorus calamus, Sparganium erectum, Alisma plantago-aquatica, Lythrum salicaria, Ranunculus lingua, Cotula coronopifolia, Salix spp., and Populus spp. The floating plant ponds (hydrophytes) contain Lemna minor, Stratiotes aloides, or Callitriche vernalis. Under-water systems (hydrophyte) contain Myriophyllum spicatum, Elodea canadensis and Ceratophyllum demersum. A majority of these plants were selected for their nutrient uptake or their Fig. 1 Schematic representation of the constructed wetland, consisting out of an intensive zone and an extensive zone and the main flow within the system; white hexagonals mark the location of flow meters, the white triangle depicts the location of climatological monitoring

known impact on the N household, others were chosen for their water household, rooting characteristics, their capacity to hygienise the substrate, for ecological purposes or finally for aesthetic appreciation of the system (landscape integration). 2.2 Analysis Conductivity was measured using an LF 537 conductivity electrode (Wissenschaftlich Technischen Werkstäten, Weilheim, Germany). pH was determined potentiometrically using an Orion model 520 pH meter (Orion, Boston, MA, USA). Total N was determined using a modified Kjeldahl procedure (Van Ranst et al. 1999). Total P content was determined using the colorimetric method of Scheel (Van Ranst et al. 1999). Two standards (25 ppm; 50 ppm) were used to calibrate absorbance in function of P concentration. The absorbance at 700 nm of samples and standards was determined using a Jenway 6400 spectrophotometer (Barloworld Scientific T/As Jenway, Felsted, United Kingdom). Chemical oxygen demand (COD) was determined photometrically using Dr. Lange standardized cuvettetests (Dr. Bruno Lange GmbH & Co, KG Düsseldorf, GE).

Intensive zone

Return

Out

Extensive zone

In

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2.3 Mass Balance Calculations The loading and discharge rates of the system were monitored by means of flow meters at inlet and outlet points. Inflow by rainfall was monitored and converted to volumetric loading rates based on the inflow surface area of the wetlands. N processed by the system between two points of time ( j and j+1) was calculated by the N loading rate minus the discharge rate, with deduction of the N mass accumulated within the system over the investigated time period. This definition implies that storage in the system is not recognized as being processed. N within the system was assessed by multiplying volume content and N concentrations for each individual basin. Processed N can be subdivided in N removed from the substrate by denitrification, N removed by plant uptake and harvest, and N removed by sludge removal. N processed ¼ N influent N effluent $N system ¼ N denitrification þN plant removal þN sludge removal X X $N system ¼ ½V i C i jþ1 ½V i C i where Vi = volume in basin/pond i; and Ci = N concentration in basin/pond i. Processed P in our monitoring was defined as the mass removed from the system by either plant harvest or sludge removal. Accumulation of P in the system between two points in time was defined as the influent load minus the effluent load, with deduction of P processed by plant or sludge removal. Pprocessed ¼ Psludge removal þ Pplant removal Psystem ¼ Pinf luent Peffluent Psludge removal Pplant removal

3 Results and Discussion 3.1 Manure Pretreatment In Flanders, industrial scale manure processing occurs mainly by (a) an initial physical separation into a liquid (85%) and a solid fraction (15%; in our case by centrifugation), (b) composting of the solid fraction with a mass reduction of 60% and (c) N reduction in the liquid fraction in an activated sludge reactor by nitrification and denitrification. The effluent of the activated sludge reactor is a nutrient poor, brackish


waste stream, which in itself is not entirely interesting for application as a fertilizer. The activated sludge reactor has removed more than 90% of the original N content and since fertilization limits are expressed in kilogram per hectare, higher loads of this nutrient poor but salt-rich effluent can be disposed off on arable land. This in turn holds risks for soil salinisation (effluent ∼12–18 mS/cm) and nitrate leaching (most of the N is converted to the more geo-mobile nitrate). Our tertiary process aims to alleviate this problem by two subsequent treatment steps to further convert the pre-treated liquid fraction into re-usable water. The first step involves a physico-chemical treatment (based on Meers et al. 2006) while nutrient and COD concentrations are reduced to below discharge criteria using a polishing step with constructed wetlands (Meers et al. 2005b). Volatilization of ammonia from ponds and lagoons for storage and treatment of animal manure is an environmental issue of concern (Aneja et al. 2000). However, as described by Meers et al. (2005a) and Vanotti et al. (2007) the N still present in the animal manure after physical separation and biological treatment (nitrification/denitrification) is mainly converted to NO 3 N , successfully reducing the risks of ammonia emissions from the ponds within the current design. 3.2 Meteorological Data Average temperatures at our site varied between 16.4– 21.5°C in summer, 9.3–14°C in autumn, 6.1–7.4°C in winter and 7.8–12.2°C in early spring (Fig. 2). These were extraordinary warm for this region. Normal temperatures in Belgium are 14.3–16.8°C in summer, 6.2–10.3°C in autumn, 2.4–3.2°C in winter and 5.2– 8.4°C in early spring. Precipitation as well was marked by unusual meteorological phenomena for Belgium, with droughts in July 2006 and April 2007 and excessive rainfall in August 2006. Normal precipitation levels in Belgium vary around 72–74 mm in summer, 64– 72 mm in autumn, 42–59 mm in winter and 46– 50 mm in early spring. 3.3 Loading Rate and Removal Efficiency Average influent concentrations into the constructed wetlands were 364±179 mg l−1 N, 58±27 mg l−1 P

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Fig. 2 Meteorological data for the test site: bars present precipitation in millimeter per month (left Y axis), line graph indicates average day temperature (right Y axis)

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and 1,739±807 mg l−1 COD. Table 1 presents the loading rates of the liquid fraction of pig manure, the corresponding masses of ingoing N, P and COD, as well as the discharge rates. These results reflect the treatment performance of the first portion of the field (1,000 m2), constructed in April 2006. The second portion of the field (3,500 m2) was constructed in autumn 2006, slowly started up in the following spring and is expected to achieve full operational status by the summer of 2007. These results show that the system is well equipped to handle high loads. In total 1,098 t of pre-treated pig manure entered the system containing 400 kg of N, 64 kg of P and 1,906 kg of COD. This implies a loading rate of 2.5 mm per day, which is higher than the 1 mm reported by Meers et al. (2005a) at lab scale. Treatment corresponded with removal efficiencies of 99.3%, 99.6% and 97.8% respectively. Plant uptake was estimated to account for 9% of the N

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removal and 31% of the P removal. Another 7% of P was recuperated from the system by sludge removal. Faulkner and Richardson (1990) reported that 5% of total N removal can be attributed to plant uptake. Lee et al. (2004) recovered 2–4% of removed N in the plants, and Newman et al. (2000) only 2.6%. Meers et al. (2005a) estimated the relative importance of N removal by plants at 7–18%, which corresponds well with our field scale monitoring results. While removal of COD and N occurred predominantly into the atmosphere by degradation and denitrification processes, P was mainly removed by precipitation onto the substrates of the various systems. This is of course not a sustainable means of removal in the long run, taking into account that once saturated the system will not be able to remove P to below discharge criteria any further. The system has therefore been adjusted to allow a more efficient sludge removal, as well as a higher P removal rate by the various plants in the

Table 1 Influent and effluent loads for the pilot scale installation from April 2006 until April 2007

Apr–Jun Jul Aug Sep Oct Nov Dec Jan Feb Mar Apr

In (m3)

N (kg)

P (kg)

200 94 138 293 0 63 0 16 28 115 151 1,098

50 16 48 137 – 30 – 7.1 28 33 51 400

40 4.8 4.0 7.2 – 0.34 – 0.16 0.78 1.7 5.4 64

COD (kg) 200 91 182 246 – 12 – 580 84 292 219 1,906

Out (m3) 0 0 160 83 0 0 0 0 351 136 0 730

N (kg)

P (kg)

COD (kg)

– – 0.6 0.5 – – – – 1.1 0.4 – 2.6

– – 0.10 0.03 – – – – 0.06 0.04 – 0.23

– – 19 6.8 – – – – 8.2 5.5 – 40

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system. Projected distribution of P removal by the adapted system will be 30–50% by plant harvesting and 50–70% by sludge removal and/or substrate regeneration, with no remaining imbalances resulting in a progressive accumulation in the sediment. To assure sustainable performance in the long run, maintenance including substrate regeneration, sludge removal and plant harvest will be an important prerequisite. Pleustophytes, such as L. minor, can be harvested continuously during the growing season in a mechanised manner, while the helophyte beds need to be harvested once per year. The optimal time for harvest of helophytes needs to be determined, yet it will most likely be selected somewhat before the end of the growing season rather than at the end or thereafter. This because of the fact that some annual helophytes, such as P. australis, tend to re-translocate

3.4 N Removal Based on the monitoring, N was considered to be the limiting component determining the tertiary system’s operational capacity. N is also the only component legally regulated by the European Union in a separate directive regarding water quality and is therefore of particular interest. No such directives currently exist for P or COD. For these reasons, N removal within the system will be discussed in further detail. Figure 3

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Effluent jun-06 jun-06 jul-06 jul-06 jul-06 aug-06 aug-06 aug-06 sep-06 sep-06 sep-06 oct-06 oct-06 oct-06 nov-06 nov-06 nov-06 dec-06 dec-06 dec-06 jan-07 jan-07 jan-07 feb-07 feb-07 feb-07 mar-07 mar-07 mar-07 apr-07 apr-07 apr-07 may-07 may-07 may-07 jun-07

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Influent Discharge criterion Effluent

0 jun-06 jun-06 jul-06 jul-06 jul-06 aug-06 aug-06 aug-06 sep-06 sep-06 sep-06 oct-06 oct-06 oct-06 nov-06 nov-06 nov-06 dec-06 dec-06 dec-06 jan-07 jan-07 jan-07 feb-07 feb-07 feb-07 mar-07 mar-07 mar-07 apr-07 apr-07 apr-07 may-07 may-07 may-07 jun-07

N (mg/l)

Fig. 3 Influent and effluent N concentrations (mg l−1) for the intensive zone (first part of the wetlands) and the extensive zone (last part of the wetlands); discharge criterion for N is presented in the second graph (15 mg l−1)

nutrients to below-ground organs towards the end of the growing season. A similar suggestion was made by Meuleman et al. (2003), who stated that removal efficiencies of N and P could be improved by harvesting the helophytes in October rather than the current harvesting practices in December.


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presents influent and effluent concentrations in both the intensive and extensive zone of the constructed wetlands. The effluent of the intensive zone is also the influent of the extensive zone, while the effluent of the extensive zone can either be re-used as water or discharged into surface water. In the current monitoring, more than 90% of the N was already removed in the intensive zone. However, the extensive zone was only constructed in fall 2006 with operational capacity to be achieved by the summer of 2007. Nevertheless, the intensive zone by itself already demonstrated extraordinary N removal capacity, on average approximately 0.89 g m−2 day−1. In some basins, designed for enhanced N removal within the system, removal rates of up to as high as 5–10 g m−2 day−1 were observed. Based on Table 1 it could appear as if the low input during autumn and winter months is due to decreased system performance during this colder period of the year. Figure 4 sheds a different light on this matter: the intensive zone was overfed in September, which led to a decreased input the following months. However, N continued to be removed from the system at a steady pace of 0.89 g m−2 day−1 on average, even during the colder months. Closer examination revealed that the removal rate was 1.22 g m−2 day−1 in the summer months, 0.75 g m−2 day−1 in autumn and winter months and again 1.13 g m−2 day−1 in early spring (March, April). This implies that, although a significant decrease in removal rate (−38%) was observed during autumn and winter, N removal never was reduced to zero meaning that the

Fig. 4 Cumulative mass of N (kg) added to the system (black bullets) and processed by the system (white bullets)

system can perform year-round. Moreover, the overall residence time within the system (both intensive and extensive zone) is designed to be sufficiently long (3– 4 months) to buffer against reduced performance in colder periods. However, considering the abnormal climatological conditions observed for this region, both in terms of temperature and precipitation, extensive monitoring will continue in subsequent years to evaluate variability in removal performance as a result of variability in climatological conditions. 3.5 P Removal Water quality standards for P are not yet regulated by a European directive, such as is the case for N. However, eutrophication with P can also result in detrimental effects on surface water ecology by causing harmful algal blooms (Ulen and Weyhenmeyer 2007). The constructed wetlands in our experiments demonstrated a very efficient removal of residual P from the pre-treated manure: regional discharge criteria are 2 mg l−1 P whereas effluent concentrations were on average below 0.3 mg l−1 P. The mechanisms by which P is removed from wastewater include sorption on substrates, storage in biomass and the formation of new sediments (Kadlec and Knight 1996). However, removal by storage on the substrate by sorption and sedimentation has a finite capacity and does not contribute to the long-term sustainable P removal (Dunne and Reddy 2005). This implies that, in order to assure sustainability of the removal performance of the system, the P loading rates need to be related to

400 350

250 y = 0.8918x

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Nitrogen (kg)

300

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the P uptake by harvestable biomass. The nutritional uptake per surface area varies greatly between aquatic plant species and different wetland designs, yet based on observed biomass productivity and plant P concentrations, annual removal rates in our system will be situated between 50 and 100 kg ha−1 P. Wetland loading rates will therefore depend on the constraints imposed by nutrient removal capacity of the plants as well as N removal rates by means of microbial processes as described in the section above. Vymazal (2007) estimates that P removal capacity by harvesting of aboveground biomass from emergent vegetation lies in the range of 10–20 g m−2 year−1, or 100–200 kg ha−1 year −1. Removal capacity as observed by Meuleman et al. (2002) more closely agrees with our findings, with removal in the range of 50–100 kg ha−1 year−1 when harvested early October. 3.6 Correspondence with Legal Criteria Table 2 compares the COD, N and P concentrations at the end of the system with the respective legal criteria for discharge into the surface water. For the discharges in August and September, the end of the

Table 2 P, N and COD concentrations at the end of the wetland system in comparison with the legal discharge criteria ; in autumn 2006 the extensive zone was constructed which means that from 2007 onwards the end of the wetland system corresponds with the end of this zone

Criteria

Ptot (mg l−1) 2

End of intensive zone 11/Aug 0.8 30/Aug 0.7 11/Sep