Phosphorus and nitrogen removal from tertiary treated urban wastewaters by a vertical flow constructed wetland

Ecological Engineering 61 (2013) 34–42

Contents lists available at ScienceDirect

Ecological Engineering journal homepage: www.elsevier.com/locate/ecoleng

Phosphorus and nitrogen removal from tertiary treated urban wastewaters by a vertical flow constructed wetland M. Martín ∗ , S. Gargallo, C. Hernández-Crespo, N. Oliver Instituto de Ingeniería del Agua y Medio Ambiente, Universitat Politècnica de València, Cno. de Vera s/n, Valencia, Spain

a r t i c l e

i n f o

Article history: Received 16 April 2013 Received in revised form 10 July 2013 Accepted 20 September 2013 Available online 19 October 2013 Keywords: Constructed wetland Eutrophication Phosphorus adsorption Nutrient removal Water treatment Vertical flow

a b s t r a c t Lake L’Albufera is a hypertrophic lake exposed to anthropic pressures. To reduce nutrient loads, a set of horizontal and vertical sub-surface flow constructed wetlands (CWs) were built to treat wastewater from a tertiary wastewater treatment plant (WWTP) before its discharge into the lake. These CWs were designed to remove nutrients, primarily total phosphorus (TP). This paper is focused on a vertical flow constructed wetland (VFCW), the primary objective of which was to remove TP by adsorption and biological uptake. Prior to construction, laboratory experiments were conducted to determine which materials and in what proportions are best suited to achieve that goal. Two different sands (types 0 and 1), as filling material, two types of clays (types 1 and 2) and two types of iron oxides (types a and b), as sorption agents, were used. The primary parameters studied were the phosphate adsorption isotherm, the filter medium hydraulic conductivity and the depletion of adsorption capacity in experimental columns. Laboratory results showed that the best mixture was formed by sand type 0 (Qmax = 2.94 mg P kg−1 ) and 10% of iron oxide type b (Qmax = 1666.67 mg P kg−1 ). Operation was established in a daily cycle with a hydraulic loading of 0.068 m d−1 . Following this procedure, a VFCW planted with reeds and with 157.9 m3 water capacity per cycle was constructed. The inflow contained 0.635 mg TP l−1 , 1.906 mg NH4 + -N l−1 , 8.9 mg TN l−1 and 20.9 mg COD l−1 . During the first year of operation, total phosphorus removal was 77.0%, ammonium 95.0%, total nitrogen 24.4% and organic matter (COD) 49.3%. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Currently, CWs are considered valuable tools to treat a wide range of wastewaters, such as industrial wastewater, mining effluents and urban and agricultural runoff (Kadlec and Wallace, 2009). In particular, VFCWs are increasingly used for specific purposes, such as sludge dewatering (Uggetti et al., 2011) or treating highstrength wastewaters (Vázquez et al., 2013). The combination of biological, physical and chemical processes that takes place in CWs allows the removal of different types of pollutants, and the removal efficiency for each parameter depends on the system design. Microbiological processes are improved by existing macrophytes, and consequently, the removal of common pollutants as biodegradable organic matter is notably effective. Nitrogen removal by nitrification–denitrification is more complex but viable with a properly designed set of CWs alternating oxic and anoxic conditions (Brix et al., 2003).

∗ Corresponding author. Tel.: +34 96 387 76 17; fax: +34 96 387 76 17. E-mail address: [email protected] (M. Martín). 0925-8574/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.ecoleng.2013.09.046

However, phosphorus removal in CWs is not as efficient as for biological oxygen demand or total suspended solids because removal processes, such as plant uptake, assimilation by microorganisms, adsorption and precipitation, are usually of little significance (USEPA, 2000). Moreover, the removal of phosphorus by plant harvesting of rooted emergent macrophytes is still a subject of debate (Kadlec and Wallace, 2009). The efficiency of phosphorus removal can be enhanced in subsurface flow constructed wetlands by using a filter media with high P binding capacity; in recent years, considerable research in this field has been focused on the study of sorbent materials and its use in VFCWs (Westholm, 2006; Vohla et al., 2011). To enable long-term phosphorus removal, many types of sorbent materials have been tested: natural materials, such as limestone, zeolites and shells, artificial products, such as Filtralite® and LECA® , and by-products, such as blast furnace slag, fly ash or dewatered sludge from water treatment facilities (Babatunde et al., 2009). Knowledge of the crucial role of phosphorus as limiting nutrient in inland waters has led to an increase in efforts to remove it from wastewaters. According to the Council Directive (91/271/EEC), the requirements for discharges from urban WWTPs to sensitive areas, which are subjected to eutrophication, are 1 or 2 mg P l−1 ,

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

depending on the population equivalent. However, it has not been determined that this target concentration is sufficient to preserve the water quality of receiving water bodies in all cases. One of the most prominent lakes in Spain is the Lake L’Albufera, located 10 km south of the city of Valencia in the east of Spain. In the 1980s, Miracle et al. (1984) reported its hypertrophic status, which continues to this day. Currently, several active approaches are being implemented to improve its state: the treatment of eutrophic waters using constructed wetlands (Martín et al., 2013), the study of sediments (Hernández-Crespo et al., 2012), runoff interception by stormwater tanks (Andrés-Doménech et al., 2012) and the improvement of waters that feed the lake. Another environmental problem in Lake L’Albufera, the decrease of high-quality freshwater inflows, is being partially balanced by a fraction of the effluent of the largest WWTP serving Valencia City, Pinedo-2 (P2WWTP) (88.5 Hm3 water treated per year). P2WWTP supplies 1 m3 s−1 of tertiary treated wastewater to L’Albufera, with a TP concentration approximately 1 mg P l−1 . In previous studies of Lake L’Albufera (TYPSA, 2004), a eutrophication water quality model was implemented, and from the simulation results, it was stated that a concentration of 1 mg P l−1 from P2WWTP might be too high to produce an enhancement of water quality, as required by the EU Water Framework Directive (WFD, 2000/60/CE). At that time, it was decided that an inflow concentration of 0.1 mg P l−1 would be a technically feasible “compromise” between the limited natural waters incoming to the lake (0.02 mg P l−1 ) and the recycled wastewater (1 mg P l−1 ). Consequently, a research project was developed with nine units of subsurface flow constructed wetland, the V30-Constructed Wetland Facility (V30-CWF), to treat part of the tertiary outflow of P2WWTP prior to its discharge in Lake L’Albufera. Six units were horizontal flow CWs, and the three others were VFCWs. The main objective was to find the maximum treatment capacity in terms of hydraulic loads and phosphorus, nitrogen and organic matter loads. This paper is focused on one of the three vertical units, that is, unit number 9: a unit specifically designed to remove phosphorus, the substrate of which was a mixture of sand as a filling material and iron oxide as a sorbent material, which was planted with common reeds (Phragmites australis). The sand was selected according to its hydraulic properties, whereas the iron oxide was selected by adsorption optimisation. In this case, special attention was paid to phosphorus adsorption processes with a sorbent agent not yet studied, and the study of the role of reeds in phosphorus removal. The primary hypothesis is that for this type of water, iron oxide is a suitable material for P adsorption and the bioassimilation is sufficiently significant to extend the lifespan of the VFCW. Following Arias and Brix (2005), before building unit 9, several laboratory studies were performed to investigate suitable materials. The results obtained in laboratory tests, the selected filter media of unit 9 and the results of unit 9 during the first year of operation are presented in this paper.

2. Materials and methods 2.1. Laboratory experiments Wastewater used for laboratory experiments was collected seven times from the outlet of P2WWTP tertiary treatment between July and November 2010. Only physicochemical variables (dissolved oxygen, pH, temperature and conductivity), total phosphorus (TP) and phosphates (PO4 3− -P) were analysed in this stage because laboratory experiments were focused on phosphorus removal. TP and PO4 3− -P were analysed using the Spectroquant®

35

Analysis System by Merck; samples were prepared for TP analysis by digestion with sulphuric acid and peroxodisulphate, and standard method ISO 6878 was used for both TP and PO4 3− -P. The TP concentration ranges between 0.69 and 1.41 mg P l−1 with a mean of 0.89 mg P l−1 ; 98% of the TP consists of PO4 3− -P. Two types of iron oxides (Fe2 O3 -hematite) and two types of clays were tested as sorbent agents; two types of sands were also tested as filling materials. The concentrations of P, Al, Fe, Ca and Mg in these materials were analysed by X-ray fluorescence; their pH values were determined according to UNE 77305:1999. The maximum phosphorus sorption capacity of the individual materials was calculated by batch laboratory experiments. Usually, this type of experiment is performed using synthetic water, but it has been stated that the results of sorption are different when real water is used (Ádám et al., 2007). In this case, the studied water was not synthetic, meaning that the range of phosphorus concentrations was conditioned by the performance of P2WWTP. This condition was an important characteristic of the experiment because it limited the range of isotherm points available and necessitated the use of an appropriate mass (usually low) of sorbent for each point of the isotherm. One litre of wastewater was stirred with a given mass of sorbent for 21 h at 24 ◦ C. Adsorbed PO4 3− -P and equilibrium P-concentration data for each material were fitted to linear expressions of the Langmuir and Freundlich models. Based on the results of the sorption isotherms, two column tests were designed. Sand was selected as the filling material and clays or iron oxides as the active agents; the first column test was performed to ensure proper drainage and establish the most appropriate proportion of sand and sorbent material to yield a suitable hydraulic conductivity (Kc ). This Kc value should be sufficiently low to flood the VFCW surface quickly, ensuring a uniform downflow across the filter media, but not sufficiently low for the drainage to be slowed. Because of the large number of possible combinations, an initial study was focused on the two individual sands (namely, sand 0 and sand 1); next, two more experiments were performed using the sand that exhibited a higher Kc (the most unfavourable) with clay 1, as its characteristics were similar to clay 2, in percentages of 5% and 10%. The last two tests were performed combining sand 1 with the worst (from the adsorption results) iron oxide and sand 0 with the best iron oxide. The experimental column was 1 m long and 0.058 m in diameter. The filter medium consisted of a 0.05 m gravel (0.5–1 cm in diameter) layer and a 0.4-m layer of the active substrate. Saturated hydraulic conductivity was determined as the minimum infiltration rate on the Horton curve (mm h−1 ), measuring the time needed for a 0.5-m water column to drop in height by intervals of 5 cm. The second column test was designed to reproduce the CW profile and operation on a laboratory scale. In accordance with the observed results, not all of the possible combinations of materials were tested; for example, if iron oxide a exhibited significant P sorption, it could be assumed that P sorption for iron oxide b it would be higher because its maximum adsorption capacity was also higher. Consequently, the number of experiments was reduced, and only three tests were performed: pure sand 1, sand 1 with 5% of clay 1 and sand 1 with 5% of iron oxide a. The column used to study phosphorus adsorption was 1.1 m long and 0.112 m in diameter; the medium consisted of, from the bottom to the top, a 0.05-m gravel layer, a 0.4 m layer of the experimental medium and a 0.05 m top layer of gravel. These columns were operated with discontinuous flow: 550 ml per cycle, four cycles per day. The hydraulic loading rate (HLR) was 0.22 m3 m−2 d−1 , and the studied hydraulic retention times (HRT) were 15, 30 and 60 min. Outflows were analysed for total suspended solids (TSS) according to the method described in the

36

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

Fig. 1. Location and scheme of the V30-Constructed Wetland Facility.

Standard Methods for Examination for Water and Wastewater (APHA, 1991) and PO4 3− -P.

2.2. V30 Constructed Wetland Facility: site description and operation The V30-CWF, which has a total surface of 2.8 ha, was formed by a set of nine units of subsurface flow constructed wetlands in parallel (Fig. 1). Each unit was fed independently; therefore, water was not mixed among the units. The units were designed to study the improvement of water quality in the effluent of P2WWTP tertiary treated wastewater, with special attention being paid to phosphorus and nitrogen removal. V30-CWF was located (39◦ 26 09 N, 0◦ 22 43 W) south of Valencia City, 4 km from the P2WWTP and located in the new Turia river bed, usually dry, which was constructed in the 1970s to divert the old Turia River (currently a city park). The projected life of the V30-CWF was only 3 yr, after which, it would be removed. The obtained results would serve as the basis for the design of a full system for P2WWTP. The study area has a typical Mediterranean climate: the annual mean precipitation is 454 mm and is concentrated mainly in autumn followed by spring. The mean temperature varies between 11.5 ◦ C (January) and 25.5 ◦ C (August), and the annual potential evapotranspiration is 866 mm and ranges between a minimum of 0.5 mm d−1 in winter and a maximum of 7.1 mm d−1 in summer. In the nine units, different combinations of the main parameters influencing phosphorus removal were studied (e.g., horizontal and vertical flow, filling material, HRT, plant species). This paper is focused on unit 9, a VFCW which is 117 m in length, 20 m in width and 0.9 m deep. This unit is divided into two parts by a 3-m-wide central path, which is used as an access path without plants. The two parts were planted with P. australis with an initial plant density of 2 plants m−2 . The VFCW profile consisted of four layers: the deepest was the lining layer, with compacted clays; above it, there was a 10-cm layer of gravel (2–3 cm in diameter); next, the active layer was formed by a mixture of sand 0 and 3% of iron oxide b; and finally, on the top, another 10-cm layer of gravel. The total amount of iron oxide was 100 tonnes. The unit 9 was operated intermittently (tidal flow operation) with one daily cycle: it was flooded with 157.9 m3 of tertiary treated wastewater supplied by means of four longitudinal pipes

(110 m long) with ten holes in each one for 45–60 min at a flow of 0.050 m3 s−1 ; the holding period was approximately 19 h, and after that, it was drained for 3.5–4 h. The number of fill and drain cycles was limited by personnel availability because none of the operations was automated. Consequently, the hydraulic load calculated on a daily basis is quite low, specifically, 0.067 m3 m−2 d−1 . The drainage system consisted of four longitudinal drainage pipes (110 m long) placed over the clays and covered by gravels; the four lines ended in a single outflow pipe.

2.3. Sample collection and analyses From October 4, 2011 to September 25, 2012, the VFCW inflow and outflow were sampled weekly, taking single samples. To determine the phosphorus concentration at different levels in the sorption media profile, three single outflow samples were taken at three different emptying times: 10 min after opening the gate (sample 9.1), 90 min later (9.2) and 210 min later (9.3). Accordingly, the first outflow sample (9.1) could be attributed to the deepest layer, where no plant interaction occurs; the second outflow sample (9.2) to an intermediate layer; and the third water sample (9.3) to the layer nearest the root zone. Water inflow and 9.3 samples were analysed using the Spectroquant® Analysis System by Merck for the following parameters: organic matter as total chemical oxygen demand (COD, ISO 15705), total nitrogen (TN, ISO 11905-1 + photometric determination of nitrate), ammonium (NH4 + -N, ISO 7150-1), nitrite (NO2 − -N, EPA 354.1), nitrate (NO3 − -N, photometric determination of a red nitro compound formed by the reaction of nitrate and a benzoic acid derivative), total phosphorus (TP) and phosphate (PO4 3− -P). Analytical quality assurance was performed by checking the measurement system with a Spectroquant® standard solution. Samples 9.1 and 9.2 were analysed for TP, phosphate and TSS once a week, but only quarterly for COD and nitrogen species. In addition, the following physicochemical water parameters were measured in situ (WTW-Multi 340i) for all the samples: water temperature, pH, conductivity and dissolved oxygen (DO). Nutrient loads from dry and wet atmospheric deposition were measured with an atmospheric total deposition sampler. Five samplings of the above-ground biomass of vegetation were collected between June 2012 and January 2013 to evaluate the

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

37

Table 1 Physical characteristics, pH and concentrations (mg g−1 dry weight (dw)) of phosphorus (P), calcium (Ca), magnesium (Mg), aluminium (Al) and iron (Fe) in the materials tested for P-adsorption capacities. Material

d10 (mm)

d60 (mm)

UC (d60 /d10 )

Bulk density (g cm−3 )

pH

P (mg g−1 dw)

Ca (mg g−1 dw)

Mg (mg g−1 dw)

Al (mg g−1 dw)

Fe (mg g−1 dw)

Sand 0 Sand 1 Clay 1 Clay 2 Iron oxide a Iron oxide b

0.076 0.103

0.472 0.448

6.2 4.3

1.6 1.5 1.3 1.0 0.8 1.2

8.77 8.66 8.62 8.57 5.37 4.94

0.22 0.22 0.22 0.22

0.71 0.71 46.48 1.97 0.43 6.15

0.60 0.60 4.43 2.29 0.60 5.19

11.66 15.04 81.04 158.33 0.72 8.64

0.70 0.70 23.97 17.02 661.38 539.47

vegetation growth and density, as well as the amount of nutrients (N and P) assimilated by plants. Only the data corresponding with summer (June–September 2012) were used to estimate the nutrient content in plants. Six representative quadrants of the VFCW were selected to delimit an area of 1 m2 where the above ground biomass was harvested. Adhered sediments were removed, and once clean, biomass was weighed in the field to determine wet density. According to official methods (MAPA, 1994), a representative subsample of all biomass was prepared for analysis in the laboratory (washed, cut, dried and milled). Analyses of nitrogen and phosphorus were also performed according to these official methods. Briefly, nitrogen was determined by a Kjeldahl digestion, followed by distillation and subsequent titration of the ammonia distilled. For phosphorus analysis, samples are ashed (450 ◦ C for 2 h), and concentrated HCl is added and mixed while heating. After cooling, the mixture is filtered, and phosphorus is determined by colorimetry according to ISO 6878. Analyses were performed in duplicate for quality assurance. Statistical analysis of the experimental results was performed using the SPSS 16 software. Populations with normal distributions were analysed using Student’s t analysis, and populations with nonnormal distributions were analysed with the Wilcoxon test. 2.4. Nutrient removal calculations The nutrient percentage removal was calculated from inflow and outflow, corrected by evapotranspiration. The potential evapotranspiration was obtained from nearby meteorological stations, but a more realistic approximation of in situ evapotranspiration could be calculated if a significant increase in output conductivities was observed. For mass balance purposes, the outflow concentration was calculated as the mean value of samples 9.1, 9.2 and 9.3. The relative concentration term, C/C0 , was used only for discussion of TP performance, where C0 is the TP input concentration and C the TP concentration in 9.1, 9.2 or 9.3. The relative importance of biomass uptake and sorption in TP removal can also be calculated with a simple mass balance: the main hypothesis is that the removal of TP in the bottom layer (represented by sample 9.1) is due only to adsorption, whereas in the intermediate (9.2) and top (9.3) layers, assimilation by plants and microorganisms is also relevant. This hypothesis would be supported if significant differences were found among samples 9.1, 9.2 and 9.3; the TP removed could subsequently be distributed between adsorbed and in biomass. The mass removed in every cycle by plants and microorganisms was calculated as the volume treated multiplied by the difference between 9.1 and the outflow concentration. Although a full mathematical model is beyond the scope of this study, an estimation of heterotrophic bacteria biomass (non-denitrifiers) was made using ˇ unek ˘ a number of the equations described in Langergraber and Sim (2005). Because the objective of this section was to gain insight about the role of bacteria in PT removal, only equations of the aerobic growth and lysis of heterotrophs were considered. Nitrifying

bacteria were not simulated because their biomass is usually much lower than that of heterotrophic bacteria. Kinetic coefficients and ˇ unek ˘ parameters at 20 ◦ C were obtained from Langergraber and Sim (2005). 3. Results and discussion 3.1. Materials characterisation Physicochemical characteristics of the materials used are shown in Table 1. The granulometry of the two sands was notably similar, with mean diameters measuring approximately 0.3 mm and uniformity coefficients being higher than the recommended value of 3.5 (Brix and Arias, 2005). Chemical composition varies greatly among the materials, and it is possible to establish differences between the groups of materials. The pH values were similar in sands and in clays, but the iron oxide values were much lower. Calcium and magnesium contents were low, and the highest values corresponded to clay 1. These components are important in the precipitation of calcium phosphates (i.e., hydroxyapatite) at high pH values, but the pH in the water was close to 7 (Table 4) and the expected precipitation was negligible. Aluminium was the major element in sands and clays, and iron predominated in iron oxides with a concentration much greater than the other elements in all of the materials. clay 2 had the highest aluminium concentration (158.33 mg g−1 ), and iron oxide a had the highest iron concentration (661.38 mg g−1 ). Aluminium in iron oxide b was 12 times greater than in iron oxide a, whereas the iron content was similar. In both iron oxides, iron and aluminium were the main components, and these elements were the most determinant in defining phosphorus adsorption capacity (Dunne et al., 2005). 3.2. Laboratory experiments The hydraulic conductivities of the two sands (Table 2), the main component of the filter media, differed greatly: the hydraulic conductivity of sand 1 was six times higher than that of sand 0. Once the iron oxides were added with percentages equal to or lesser than 10% in dry weight, the Kc was reduced by half in both cases. This is a favourable result, because one of the objectives was the filling of the VFCW as homogeneous as possible, but without impair the emptying time. Table 2 Hydraulic conductivity (mm h−1 ) in the experimental medium.

a b c d e f

Active substrates

Hydraulic conductivity (mm h−1 )

Sand 0 Sand 1 Sand 1 and 5% clay 1 Sand 1 and 10% clay 1 Sand 1 and 5% iron oxide a Sand 0 and 10% iron oxide b

189.1 1209.5 1019.8 862.9 546.1 84.8

38

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

Table 3 Freundlich and Langmuir adsorption constants and correlation coefficients. Langmuir isotherm −1

Qmax (mg P kg Sand 0 Sand 1 Clay 1 Clay 2 Iron oxide a Iron oxide b

Freundlich isotherm

)

b (l mg

2.94 3.13 30.86 42.02 370.37 1666.67

−1

R

KF (l mg−1 )

n

R2

0.79 0.92 0.95 0.95 0.99 0.84

2.82 2.45 31.64 25.38 387.69 1392.83

3.23 3.45 3.70 2.17 5.26 2.38

0.87 0.73 0.79 0.97 0.90 0.90

2

)

8.5 3.65 12.46 1.56 33.75 6.00

The adsorption batch experiments revealed that the three groups of materials had very different maximum adsorption capacities. The adjustments that were carried out indicated good fittings for both the Langmuir and Freundlich isotherms with all linear correlation coefficients higher than 0.70 (Table 3). From Langmuir isotherms, the sands had the lowest adsorption capacities with values of approximately 3 mg P kg−1 , which are low compared to that obtained by Del Bubba et al. (2003) in Danish sands; the sands were followed by clays, with capacities approximately ten times higher. Iron oxides presented the highest adsorption capacities, which were notably different for each type. The capacity of iron oxide b was four times greater than that of iron oxide a because of different chemical compositions because the iron concentrations were similar in both but the aluminium concentration was higher in iron oxide b. According to Bolan and Barrow (1984), in general, aluminium oxides’ capacity for phosphorus adsorption is higher than that of iron oxides. Owing to the significant sorption difference between both materials, a column sorption experiment for iron oxide b was considered unnecessary. The laboratory column sorption experiments confirmed that the sands exhibited a limited P removal capacity, and the assayed clays were not suitable sorption materials. On the contrary, in columns where iron oxides were present (5% in dry weight), the PO4 3− -P concentrations were always lower than 0.01 mg P l−1 for a total of 80 cycles and an HRT of 15 min. With a mixture of sand and clays, a significant difference was observed: with this mixed media, 35 cycles were sufficient to obtain a C/C0 relationship higher than 0.7. The sands were exhausted soonest: only 23 cycles were required to obtain the same relationship. Based on the experimental studies, sand 0, due to its hydraulic conductivity, and iron oxide b, which exhibited a higher phosphorus adsorption capacity, were eventually selected. The percentages of sand (97%) and iron oxide b (3%) in VFCW were decided for economic reasons but also because with 3% iron oxide, the expected life span was close to the three-yr duration of the project. The total dry mass of the VFCW was 3300 tonnes.

3.3. Water treatment performance results In 1 yr, a total of 241 cycles was performed in the VFCW, treating a total volume of 38,061 m3 from the outflow of P2WWTP. The mean annual hydraulic load was 16.3 m yr−1 (0.044 m3 m−2 d−1 ). The potential evapotranspiration was 1166 mm yr−1 and the total annual precipitation was 285 mm yr−1 . The difference between evapotranspiration and precipitation increases the output conductivity by 7%, so removal efficiencies were modified accordingly. The mean inflow concentrations (outflow from P2WWTP) were typical for tertiary treated wastewaters (Table 4), and they satisfied, as an annual mean, the requirements of official guidelines for sensitive areas. The mean TP concentration for the studied period was 0.635 mg P l−1 , which was lower than the regulated guideline of 1.0 mg P l−1 , but higher than the desirable output value of 0.1 mg P l−1 . The mean TN concentration was 8.9 mg N l−1 , of which nitrate was the predominant species (64%). Throughout the year, different removal efficiencies were observed in P2WWTP, specifically for phosphorus and nitrogen species; and extreme variability was observed in TP input concentrations, ranging between 0.082 and 2.200 mg P l−1 , and NH4 + , ranging between 0.016 and 9.160 mg N l−1 . The TP and TN wastewater loads were 10.33 g P m−2 yr−1 and 144.76 g N m−2 yr−1 , whereas the atmospheric loads were only 0.04 g P m−2 yr−1 and 0.79 g N m−2 yr−1 . The mean output values of TP (Table 4 and Fig. 2a) were below 0.2 mg P l−1 until the last month. The best result was obtained in sample 9.3, close to the surface, and the worst in sample 9.1, at the bottom of the VFCW. This difference was statistically significant (p < 0.05) in winter, spring and summer, whereas differences between 9.2 and 9.3 were only significant in spring. This means that the TP concentration in the upper part of the VFCW was influenced by plants and attached microorganisms in the growing season. In September 2012, the output concentrations suddenly increased to approximately 0.6 mg P l−1 in 9.1 and 0.4 mg P l−1 in 9.3. This increase coincides with a period of high input TP concentrations and suggests that filter breakthrough may have been

Table 4 Mean ± S.E. (n) during the first year of operation for the selected water chemistry variables in the inflow and outflow for VFCW. Input (mean ± S.E. (n))

Variable

Output (mean ± S.E. (n)) 9.1

◦

Temperature DO Conductivity pH TSS TP PO4 3− -P TN NH4 + -N NO2 − -N NO3 − -N

C mg l−1 ␮S cm−1 s.u. mg l−1 mg l−1 mg l−1 mg l−1 mg l−1 mg l−1 mg l−1

Total COD

mg l−1

21.2 3.3 1688 7.13 2.32 0.635 0.521 8.9 1.906 0.351 5.69

± ± ± ± ± ± ± ± ± ± ±

0.7 (47) 0.2 (47) 12 (46) 0.03 (46) 0.33 (47) 0.087 (47) 0.080 (47) 0.5 (47) 0.282 (47) 0.115 (47) 0.44 (47)

20.9 ± 0.6 (47)

19.9 6.6 1739 7.06 6.53 0.174 0.137 7.6 0.198 0.036 6.72

9.2 ± ± ± ± ± ± ± ± ± ± ±

0.9 (46) 0.2 (46) 21 (46) 0.03 (46) 0.78 (46) 0.016 (46) 0.016 (46) 1.1 (3) 0.191 (4) 0.009 (3) 0.92 (3)

12.8 ± 1.6 (3)

20.0 6.9 1755 7.15 1.05 0.142 0.122 7.7 0.290 0.046 6.53

9.3 ± ± ± ± ± ± ± ± ± ± ±

0.9 (46) 0.2 (46) 20 (46) 0.03 (46) 0.09 (45) 0.013 (46) 0.013 (46) 0.8 (4) 0.240 (5) 0.015 (4) 0.74 (4)

13.3 ± 0.6 (4)

20.2 7.2 1765 7.16 2.67 0.139 0.117 7.3 0.111 0.039 6.72

± ± ± ± ± ± ± ± ± ± ±

0.9 (46) 0.20 (46) 20 (46) 0.03 (45) 1.01 (46) 0.012 (46) 0.012 (46) 0.3 (46) 0.039 (46) 0.008 (45) 0.36 (46)

11.7 ± 0.3 (46)

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

39

Fig. 2. Time series of the input–output water quality in the VFCW. (a) TP (input concentration in main y-axis and output concentrations in secondary y-axis), (b) DO, (c) TN and (d) NH4 + -N (input concentration in main y-axis and output concentrations in secondary y-axis).

reached. However, the relative concentration C/C0 (Fig. 3) in this final period was not very different from that in previous months. The C/C0 relationship highly depends on C0 : when tertiary treatment worked better than usual, the TP input to V30-CWF was sometimes below 0.1 mg P l−1 and the output was higher than 0.1 mg P l−1 (December–March period), indicating filter washout. For C0 higher than 0.5 mg P l−1 , C/C0 was approximately 0.2–0.3 or even lower. The experimental period ends with a value C/C0 of 0.5, but it cannot be determined if this finding indicates the start of the exhaustion of sorption capacity.

Fig. 3. Time series of TP relative concentration C/C0 (main y-axis) and TP input concentration, C0 , (secondary y-axis). The C value is the mean of TP in 9.1, 9.2 and 9.3.

The mean removal efficiencies for TP and PO4 3− -P are close to 75% (Table 5). This is an excellent result compared with other pilotscale studies; Prochaska and Zouboulis (2006) found a maximum removal of 48.5% with dolomite, and Zhao et al. (2009) reported an 88.6% removal of PO4 3− -P with dewatered alum sludge cakes. However, performance was worse than expected from laboratory columns, where the output phosphate concentrations were always lower than 0.005 mg P l−1 . This is a consequence of the scale change and the real performance of a real system: factors, such as preferential flows in water distribution, both in filling and draining, or the mixing of sand and iron in the construction stage, are not as controlled as in laboratory experiments. Therefore, the theoretical local equilibrium with iron oxide is not completely reached in any region of the filter media. The removal of PO4 3− -P and TP was satisfactory: the output concentrations were close to the objective, which should meet nutrient requirements for discharges to sensitive zones when the WWTP does not work properly, as occurs in summer (Fig. 2a). The removal efficiency of TP depends strongly on input concentration (Fig. 4). For input concentrations higher than 0.5 mg P l−1 , removal is higher than 70%, but at lower input concentrations the efficiency drops dramatically. Dissolved oxygen concentration increased notably in every cycle, from a mean value of 3.3 to 7.2 mg l−1 (Table 4 and Fig. 2b). This increase of oxygen is a common feature of filling-draining cycles, but, contrary to what happens when water has a high oxygen demand (CBOD and/or NBOD), it was not depleted in this case. This result was excellent for the receiving waters, showing that this type of CW contributes to the enhancement of aquatic ecosystems in more than just nutrient removal.

40

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

Table 5 Removal efficiencies (%) obtained in each depth in unit 9. TP

PO4 3− -P

TN

NH4 + -N

NO2 − -N

NO3 − -N

COD

9.1 9.2 9.3

73.5 77.3 78.9

74.7 76.4 78.5

– – 24.4

– – 95.0

– – 90.6

– – 17.8

– – 49.3

Total

77.0

75.2

24.4

95.0

90.6

17.8

49.3

At the end of the first year of operation, a total of 83.24 kg N and 18.30 kg P was removed from treated wastewater. This value was far from the total sorption maximum capacity, estimated at 176.4 kg P. If the TP was removed only by adsorption, 10.4% of the VFCW could be considered exhausted and the sorption capacity would still be high. Additionally, in any CW focused on nutrient removal, plant growth must be studied. The role of plants in nutrient removal can be estimated from plant biomass density and nutrient content. At the end of September, the P. australis population was well-developed (mean height of 2 m with an 80% cover); its above-ground biomass density was 1.28 kg dry weight m−2 with nutrient contents of 13.37 g

N kg−1 and 1.28 g P kg−1 . The nutrient contents were in the low range of those reported in the literature (Tanner, 1996). The TN and TP content in above-ground biomass was 34.04 kg N and 3.26 kg P. If the roots are considered, and taking a typical relationship between nutrients in above-ground and below-ground plant components of 1.33:1 for N and 1:1 for P (Tanner, 1996), the total amount of nitrogen and phosphorus in plants was 79.31 kg N and 6.52 kg P. Plant growth accounts for 35.6% of the TP and 95.3% of the TN removed. These are high percentages, leading to the conclusion that the plants play a significant role in nutrient removal when input concentrations are low (Vymazal and Kropfelova, 2008). Applying the mass balance described in Section 2.4, it was possible to distribute the TP removed between adsorbed and in biomass (Fig. 5). At the end of the study period, the TP in biomass, 6.20 kg P, was very close to the 6.52 kg P obtained from experimental vegetation data. The mass balance results confirm that concentration differences among filter layers were due to biomass assimilation. Considering that the organic matter load was low, 340 g COD m−2 yr−1 , microbial biomass was expected to have a minimal impact on TP retention. Tietz et al. (2007) reported values of bacterial biomass from 50 to 300 ␮g C g−1 dw in the first 10 cm of a vertical flow constructed wetland with an organic load of 7300 g COD m−2 yr−1 . From these authors’ data, a mean value of 29 ␮g C g−1 dw can be estimated at the totality of filter. Assuming that all the COD removed (mean value of 9.2 mg l−1 ) was biodegraded and steady state was reached, the biomass of aerobic heterotrophic bacteria in the VFCW was calculated as explained in Section 2.4. The mass in terms of COD is, on 17.8 kg COD and the corresponding mass of carbon represented by the bacteria is 6.67 kg C or 2.01 ␮g C g−1 dw. This value is 14 times lower that obtained by Tietz et al. and is consistent with the organic loads. Consequently, with a composition of 0.02 mg P mgCOD,BM −1 and 0.07 mg N mgCOD,BM −1 ˇ unek, ˘ (Langergraber and Sim 2005), approximately 0.36 kg P and 1.26 kg N were removed from the water by bacteria. The phosphorus and nitrogen in bacteria only accounts for less than 2% of the nutrients removed. Only when organic loads are very high can the TP removed by microorganisms reach a significant percentage (Kumar et al., 2011).

Fig. 4. TP percentage removal versus TP input concentration.

Fig. 5. Cumulative removal mass of TP, total, by biomass, and by sorption in the main y-axis. TP input concentration is represented in the secondary y-axis.

Removal of ammonia (Table 5 and Fig. 2d) and nitrites was quite high; the nitrification process with high values of dissolved oxygen could explain part of this fact. However, while a net increase of nitrates was observed in the output, some degree of DIN removal was achieved because the increase of nitrates was lower than the sum of ammonia and nitrified nitrites. Total nitrogen removal was low (Table 5), measuring only 24.4%, but taking into account that the VFCW is not specifically designed to remove TN, this finding is a satisfactory collateral result (Fig. 2c). The reduction by half of the COD input concentration was another collateral positive consequence of the treatment, which was unexpected because COD was already low after tertiary treatment. The biodegradability of COD can be estimated from the available output data from P2WWTP in 2011. The annual mean COD concentration was 20 mg l−1 (similar to what was measured in this study as input to V30-CWF in 2012) whereas BOD5 was between 2 and 5 mg l−1 . Consequently, the COD was poorly biodegradable, with a BOD5 /COD relationship of 0.15, even lower than the 0.33 value obtained by Ayaz (2008) in a similar study. In spite of this, the attached bacteria in plant roots and porous media were able to significantly reduce COD, with an HRT lower than one day. The removal efficiency achieved in the cell (49.3%) was very close to the 50% removal presented by Ayaz (2008). 3.4. Sorption, biomass and nutrients removal

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

41

biomass (sample 9.1), the filter is ineffective (C/C0 = 1) when TP input concentration is lower than 0.13 mg P l−1 , but when biomass is included (sample 9.3), this value decreases to 0.08 mg P l−1 . The presence of biomass, mainly plants, increases the range of validity of this type of VFCW. Plant uptake of phosphorus could increase the designed life span of 3.5 yr, considering only sorption and one daily cycle per year, to 4.9 yr. However, this value could increase further because the plant density could be expected to rise in subsequent years.

4. Conclusions

Fig. 6. (a) Correlations between TP input concentration and TP mass removed per cubic metre of treated wastewater, by biomass and by sorption. (b) Percentage contribution of biomass and sorption to TP removal, depending on the concentration of TP input.

The relationship between TP input concentration and TP concentration removed is linear for both the portion assimilated by biomass and the adsorbed portion (Fig. 6a). Conversely, the percentage removed by sorption or biomass versus the TP input concentration is very illustrative (Fig. 6b). When the input concentration is higher than 0.50 mg P l−1 , the plants were only able to remove between 20 and 40% of TP, and adsorption is the main removal mechanism. At input concentrations lower than 0.25 mg P l−1 , the adsorption was weak and biomass became more important. The role of the biomass can also be observed in Fig. 7. Without

CWs are only rarely used to treat tertiary-treated wastewater from a WWTP. The vertical sub-surface flow CW at the V30-CWF, planted with reeds and containing a filter medium that consists of sand and iron oxide, is appropriate to reduce total phosphorus from this type of water. TP removal efficiency is higher than 75%, and output concentrations are approximately 0.1–0.2 mg P l−1 . It has been demonstrated that the reeds play a crucial role in the CW’s life expectancy because at the end of the first year of operation, 35.6% of the TP removed is in plant biomass. The importance of the reeds also can be observed in the vertical profiles of TP because a significant difference has been observed among samples obtained at different levels. Plant uptake at low concentrations of TP has been shown to be important, while bacteria have little significance as a sink of TP, owing to the low biodegradable COD load. This type of VFCW is more efficient after tertiary treatment if the input TP is higher than 0.5 mg P l−1 . Additional benefits of the VFCW include an increase in DO, the nearly complete nitrification of ammonia, a removal efficiency of 24.4% in TN due to plant uptake (95.3%) and the removal of nearly 50% in slowly biodegradable COD. Low carbon availability limits the growth of microorganisms; therefore, an increase of organic matter could enhance the nutrients retained in biomass and also promote denitrification. The management of vegetation could contribute to carbon loads to the system. After 1 yr of operation, an increase of the number of cycles to 1.5 d−1 would be possible, thereby maintaining the life span previously calculated due to the growth of reeds. The hydraulic load would increase from 0.067 to 0.1 m3 m−2 d−1 , and consequently, the total surface necessary to treat 1 m3 s−1 from P2WWTP would be 86.4 ha.

Acknowledgements We acknowledge the support of Aguas de las Cuencas Mediterráneas (ACUAMED, MMARM), and especially to V. Botella, project manager, to this research, and the attendance of ROMYMAR S.A. in the daily management of V30-CWF.

References

Fig. 7. Correlations between C/C0 and C, for TP, in 9.1 and 9.3.

Ádám, K., Krogstadb, T., Vrale, L., Søvikc, A.K., Jenssen, P.D., 2007. Phosphorus retention in the filter materials shellsand and Filtralite P®- Batch and column experiment with synthetic P solution and secondary wastewater. Ecol. Eng. 29, 200–208. Andrés-Doménech, I., Montanari, A., Marco, J.B., 2012. Efficiency of storm detention tanks for urban drainage systems under climate variability. J. Water Resour. Plann. Manag. 138 (1), 36–46 (ASCE). APHA, 1991. Standard Methods for the Examination of Water and Wastewater, 17th ed. American Publish Health Association, Washington, DC, USA. Arias, C.A., Brix, H., 2005. Phosphorus removal in constructed wetlands: can suitable alternative media be identified? Water Sci. Technol. 51 (9), 267–273. Ayaz, S.C¸., 2008. Post-treatment and reuse of tertiary treated wastewater by constructed wetlands. Desalination 226, 249–255.

42

M. Martín et al. / Ecological Engineering 61 (2013) 34–42

Babatunde, A.O., Zhao, Y.Q., Burke, A.M., Morris, M.A., Hanrahan, J.P., 2009. Characterization of aluminium-based water treatment residual for potential phosphorus removal in engineered wetlands. Environ. Pollut. 157, 2830–2836. Bolan, N.S., Barrow, N.J., 1984. Modelling the effect of adsorption of phosphate and other anions on the surface charge of variable charge oxides. J. Soil Sci. 35 (2), 273–281. Brix, H., Arias, C.A., Johansen, N., 2003. Experiments in a two-stage constructed wetland system: nitrification capacity and effects of recycling of nitrogen removal. In: Wetlands-Nutrients, Metals and Mass Cycling. Backhuys Publishers, Leiden, The Netherlands, pp. 237–258. Brix, H., Arias, C.A., 2005. The use of vertical flow constructed wetlands for onsite treatment of domestic wastewater: New Danish guidelines. Ecol. Eng. 25, 491–500. Del Bubba, M., Arias, C.A., Brix, C., 2003. Phosphorus adsorption maximum of sands for use as media in subsurface flow constructed reed beds as measured by the Langmuir isotherm. Water Res. 37, 3390–3400. Directive 91/271/EC of 21 May 1991 concerning urban waste water treatment. DO L 135 de 30.5.1991, p. 40. Dunne, E.J., Culleton, N., O’Donovan, G., Harrington, R., Daly, K., 2005. Phosphorus retention and sorption by constructed wetland soils in Southeast Ireland. Water Res. 39, 4355–4362. ˜ Hernández-Crespo, C., Martín, M., Ferris, M., Onate, M., 2012. Measurement of acid volatile sulphide and simultaneous extracted metals in sediment from Lake Albufera (Valencia, Spain). Soil Sedim. Contam. Int. J. 21 (2), 176–191. Kadlec, R.H., Wallace, S.D., 2009. Treatment Wetlands, 2nd ed. CRC Press, Boca Raton, FL. Kumar, J.L.G., Zhao, Y.Q., Babatunde, A.O., 2011. Process-based modelling of phosphorus removal in a novel constructed wetland system using dewatered alum-sludge as substrate. Water Sci. Technol. 64 (3), 774–780. ˇ unek, ˘ Langergraber, G., Sim J., 2005. Modelling variable saturated water flow and multi component reactive transport in constructed wetland. Vadose Zone J. 4, 924–938. MAPA (Ministerio de Agricultura, Pesca y Alimentación), 1994. Métodos Oficiales ˜ de Análisis. Dirección General de Política Alimentaria, Madrid, Espana., ISBN 8449100046. Martín, M., Oliver, N., Hernández-Crespo, C., Gargallo, S., Regidor, M.C., 2013. The use of free water surface constructed wetland to treat the eutrophicated waters of lake L’Albufera de Valencia (Spain). Ecol. Eng. 50, 52–61.

Miracle, M.R., García, M.P., Vicente, E., 1984. Heterogeneidad espacial de las comunidades fitoplanctónicas de la Albufera de Valencia. Limnetica 1 (1), 20–31. Prochaska, C.A., Zouboulis, A.I., 2006. Removal of phosphates by pilot vertical-flow constructed wetland using a mixture of sand and dolomite as substrate. Ecol. Eng. 26, 293–303. Tanner, C.C., 1996. Plants for constructed wetland treatment systems—A comparison of the growth and nutrient uptake of eight emergent species. Ecol. Eng. 7, 59–83. Tietz, A., Kirschner, A., Langergraber, G., Sleytr, K., Haberl, R., 2007. Characterization of microbial biocoenosis in vertical subsurface flow constructed wetlands. Sci. Total Environ. 380, 163–172. Técnica y Proyectos S.A. (TYPSA), 2004. Estudio para el desarrollo sostenible de l’ Albufera de Valencia. Confederación Hidrográfica del Júcar. Ministerio de Medio Ambiente, http://www2.chj.gob.es/albufera/index.html (7/10/2013). Uggetti, E., Ferrer, I., Molist, J., García, J., 2011. Technical, economic and environmental assessment of sludge treatment wetlands. Water Res. 45, 573–582. UNE 77305:1999. Soil quality. Determination of pH. U. S. Environmental Protection Agency, 2000. Constructed Wetlands Treatment of Municipal Wastewaters. Office of Research and Development, Cincinatti, Ohio (EPA/625/R-99/010). Vázquez, M.A., de la Varga, D., Plana, R., Soto, M., 2013. Vertical flow constructed wetland treating high strength wastewater form swine slurry composting. Ecol. Eng. 50, 37–43. Vohla, C., Kõiv, M., Bavor, H.J., Chazarnec, F., Mander, U., 2011. Filter material for phosphorus removal from wastewater in treatment wetlands—A review. Ecol. Eng. 37, 70–89. Vymazal, J., Kropfelova, L., 2008. Wastewater Treatment in Constructed Wetlands with Horizontal Sub-Surface Flow. Springer Verlag. Water Framework Directive, 2000/60/EC, of the European Parliament and of the Council establishing a framework for the Community action in the field of water, policy. OJ L 327 on 22 December 2000. Westholm, L.J., 2006. Substrates for phosphorus removal – potential benefits for on-site wastewater treatment? Water Res. 40, 23–36. Zhao, Y.Q., Zhao, X.H., Babatunde, A.O., 2009. Use of dewatered alum sludge as main substrate in treatment reed bed receiving agricultural wastewater: long-term trial. Bioresour. Technol. 100, 644–648.