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Nutrient and metal removal in a constructed wetland for wastewater treatment from a metallurgic industry
e c o l o g i c a l e n g i n e e r i n g 2 6 ( 2 0 0 6 ) 341–347
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Nutrient and metal removal in a constructed wetland for wastewater treatment from a metallurgic industry a ˜ a , H. Hadad a , G. Sanchez ´ M.A. Maine a,∗ , N. Sune , C. Bonetto b a
Qu´ımica Anal´ıtica, Facultad de Ingenier´ıa Qu´ımica, Universidad Nacional del Litoral, Santiago del Estero 2829 (3000), Santa Fe, Argentina b Instituto de Limnolog´ıa “Dr. Ringuelet”, Av. Calchaqui km 23.5 (1888), Florencio Varela, Buenos Aires, Argentina
a r t i c l e
i n f o
a b s t r a c t
Article history:
This contribution summarizes the nutrient and metal removal of a free water surface con-
Received 1 April 2005
structed wetland, compares it with the previous small-scale prototype and discusses the
Received in revised form 9
observed differences. Several locally available macrophyte species were transplanted into
December 2005
the wetland. Eichhornia crassipes (water hyacinth) showed a fast growth and it soon became
Accepted 28 December 2005
dominant, attaining 80% cover of the wetland surface. Typha domingensis (cattail) and Panicum elephantipes (elephant panicgrass) developed as accompanying species attaining 14 and 4% cover. The wetland removed 86% of Cr and 67% of Ni. Zn concentrations were below
Keywords:
50 g l−1 in most samplings. The FeS precipitation probably caused the high retention of
Metal
Fe (95%). The outcoming water was anoxic in most samplings. Phosphate and ammonium
Nutrient
were not retained within the wetland while 70% and 60% of the incoming nitrate and nitrite
Removal
were removed. Large denitrification losses are suggested. Cr, Ni and Zn were retained by the
Wastewater
macrophytes in the larger wetland and in sediment in the small-scale one. Differences in
Wetland
the retention mechanism of the two wetlands are discussed. © 2006 Elsevier B.V. All rights reserved.
1.
Introduction
The construction of artificial wetlands for wastewater treatment is now a widely accepted and increasingly common treatment alternative. Constructed wetlands were initially utilized for nutrient removal in residential and municipal sewage, storm water and agricultural runoff displaying a wide range of removal efficiencies. The accelerating industrialization in developing countries with an enormous consumption of metals constitutes an environmental contamination hazard. The application of wetlands for industrial wastewater treatment is a promising alternative. However, current experience in Argentina remains largely unreported. Conditions are favourable since there is a large availability of marginal land around most cities with a low population density. The central and northern areas of the country have mild winters, allow-
∗
Corresponding author. Tel.: +54 342 4571164; fax: +54 342 4571162. E-mail address: amaine@fiqus.unl.edu.ar (M.A. Maine). 0925-8574/$ – see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2005.12.004
ing extended growing periods for plants. Macrophytes are the main biological component of wetlands. They not only assimilate pollutants directly into their tissues, but they also act as catalysts for purification reactions by increasing the environment diversity in the root zone and promoting a variety of chemical and biochemical reactions which enhance purification (Jenssen et al., 1993). Water hyacinth (Eichhornia crassipes) is one of the most commonly used plants in constructed wetlands because of its fast growth rate and large uptake of nutrients and contaminants. It attains dense stands in the floodplain wetlands of the Middle Parana´ River close to the study site. Bahco S.A., a metallurgic factory, constructed a small-scale experimental free water surface wetland to assess the feasibility of treating wastewater from the Santo Tome´ (Argentina) tool manufacturing plant. The wetland removed 81%, 66% and
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59% of the incoming Cr, Ni and Zn, respectively, and 84% and 75% of the inorganic N and soluble reactive P (SRP) from the incoming wastewater (Maine et al., 2005). Because of the high removal efficiency attained in the small-scale experimental wetland, a large-scale facility was constructed for wastewater treatment of the whole factory. The present contribution assesses the removal efficiency of the larger wetland, compares it with the previous small-scale prototype and discusses the observed differences.
2.
Materials and methods
2.1. Design of the small-scale and the larger constructed wetlands The wetlands were constructed at Bahco Argentina metal´ Argentina (S31◦ 40 01.9 ; lurgic plant, located in Santo Tome, W60◦ 47 06.9 ). The small-scale experimental wetland was 6 m length, 3 m wide and 0.4 m deep (Fig. 1A). A polyethylene impermeable film was placed at the bottom and a soil layer of 30 cm was added. The influent entered the wetland through a PVC tube (diameter: 63 mm) with a perpendicular drip dispersion tube with aligned holes to produce a laminar flow. The inflow discharge was 1000 l d−1 and the approximate hydraulic residence time was 7 days. Three floating (Pistia stratiotes, E. crassipes and Salvinia rotundifolia) and eight emergent (Cyperus alternifolius, P. elephantipes, Thalia geniculata, Polygonum punctatum, Pontederia cordata, Pontederia rotundifolia, Typha domingensis and Aechmea distichantia) macrophytes were transplanted (Maine et al., 2005). The larger free water surface wetland was 50 m length by 40 m wide and 0.5–0.8 m deep, with a central baffle (Fig. 1B). The baffle doubled the flowpath and resulted in a 5:1 length–width ratio. The wetland received wastewater through a PVC pipe provided with a perpendicular distribution pipe with holes at regular distances in order to allow uniform distribution of flow. Wastewater discharge was approximately 100 m3 d−1 and the hydraulic residence time ranged from 7
Fig. 1 – Scheme of the (A) small-scale wetland and (B) large-scale wetland.
to 12 days. The wetland was rendered impermeable by means of a bentonite layer covered with a 1 m-layer of the excavated soil. Several locally available macrophytes were transplanted into the wetland, being E. crassipes, T. domingensis and P. cordata those of the greatest cover. After crossing the wetland, the effluent followed an excavated channel to a 1.5 ha pond. Both wastewater from the industrial processes and sewage from the factory were treated together. It was expected that high nutrient concentrations could increase the toxicity tolerance of the macrophytes (Manios et al., 2003). Effluents reached the wetland after a primary treatment. During the first 3 months of operation only sewage entered the wetland. Later, industrial waste plus sewage were treated.
2.2.
Sampling
Twenty-four water samplings were performed at the inlet and outlet of the larger free water surface wetland from November 2002 through February 2004. Samples were taken monthly until March 2003, and every 2 weeks afterwards. Sediment total P (TP), Ni, Cr and Zn concentrations were determined twice in a month at the inlet and outlet of the wetland. Sediment samples were collected using a 4 cm-diameter PVC corer. All the samples were transported to the laboratory at 4 ◦ C. In order to estimate the biomass, emergent and floating macrophyte stands were sampled with a 0.50 m × 0.50 m sampler following Vesk and Allaway (1997). Four replicates were taken. Macrophytes were then harvested and sorted by species at the laboratory, washed, separated between aboveground (shoots and leaves) and belowground (roots and rhizomes) parts, and dried at 105 ◦ C until constant weight was reached (APHA, 1998). Plant cover was estimated measuring the area occupied by each stand within the wetland.
2.3.
Chemical analysis
Conductivity was measured with a 33 model YSI conductimeter, O2 with a Horiba OM-14 portable meter and pH with an Orion pH-meter. Water samples were filtered through Millipore membrane filters (0.45 m) for P and N determinations. Chemical analysis was performed following APHA (1998); NO2 − was determined by coupling diazotation followed by a colorimetric technique, NH4 + and NO3 − by potentiometry (Orion ion selective electrodes, sensitivity: 0.01 mg l−1 of N, reproducibility: ±2%). Soluble reactive phosphate (SRP) was determined by the colorimetric molybdenum blue method (Murphy and Riley, 1962). Ca2+ and Mg2+ were determined by EDTA titration. Na+ and K+ were determined by flame emission photometry. Alkalinity (CO3 2− and HCO3 − ) was measured by HCl titration. Cl− was determined by the argentometric method. SO4 2− was assessed by turbidimetry. COD was determined by the open reflux method and BOD by the 5-day BOD test (APHA, 1998). Fe, Cr, Ni and Zn concentrations were determined in water samples by atomic absorption spectrometry (by flame or electrothermal atomization, according to the sample concentration, Perkin-Elmer 5000), following APHA (1998). Total phosphorous (TP) in sediment was determined after an acid digestion with an HClO4 /HNO3 /HCl (7:5:2) mixture followed by SRP determination in the digested samples (Murphy
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Table 1 – Measured variables in the constructed wetland during the study period Inlet
Parameter
◦
Temperature ( C) Conductivity (mS cm−1 ) O2 (mg l−1 ) pH Alkalinity (mg l−1 CaCO3 ) Ca (mg l−1 ) SO4 2− (mg l−1 ) N–NO3 − (mg l−1 ) N–NO2 − (mg l−1 ) N–NH4 + (mg l−1 ) Inorganic N (mg l−1 ) Fe (mg l−1 ) Cr (g l−1 ) Ni (g l−1 ) Zn (g l−1 ) SRP (mg l−1 ) COD (mg l−1 ) BOD (mg l−1 )
Outlet surface
Mean
Range
Mean
Range
16.9 2.9 2.4 8.7 463 156 957 16 0.79 1.9 19 13.7 22 17 <0.05 0.20 204 41
8.0–27 0.4–8.5 0–7.1 7.2–10.8 205–1187 27.0–651 98.1–2506 1.6–68 0.021–3.4 0.1–12 1.8–74 0.16–74 3.3–150 6.1–60 <0.05 0.003–0.51 22–430 7–89
16.9 1.3 1.5 7.2 237 43 395 3.0 0.15 1.6 5.0 0.38 3.6 9.0 <0.05 0.19 39 15
8.0–27 0.47–2.6 0–5.6 6.9–8.1 95–475 22.3–61.2 159–855 0.8–7.9 0.01–0.99 0.12–6.7 1.2–10 0.05–1.2 2.8–5.3 3.9–27 <0.05 0.02–0.43 15–103 5–32
and Riley, 1962). Cr, Ni and Zn sediment concentrations were determined in the digests by atomic absorption spectrometry (Perkin-Elmer 5000). Above and belowground macrophyte tissues were ground for P and metal determination. TP, Cr, Ni and Zn in plant tissues were determined in the same way as in sediment samples. Chlorophyll was extracted with acetone for 48 h under dark and cold conditions (3–5 ◦ C). The transmittance percentage of the extracts was read to 645 and 665 nm using a UV–vis spectrophotometer following Vollenweider (1974) in order to calculate chlorophyll a concentration. A nearby undisturbed floodplain wetland of the Parana´ river was also sampled to compare plant height, chlorophyll and nutrient contents in plants with those of the wastewater wetland.
2.4.
Statistical analysis
Statistical significance between inlet and outlet water concentrations was assessed using a paired-sample comparison test (p < 0.05). Differences in the concentrations of TP and metals between inlet and outlet sediment and between initial and final concentrations were determined by a mean comparison test (p < 0.05). Calculations were carried out with the Statgraphics Plus 3.0 program.
3.
Results and discussion
Table 1 summarizes the variables measured at the inlet and outlet of the wetland and the estimated removal efficiency. Comparing the concentration of variables measured at the inlet and the outlet in the different sampling, there are significantly statistical differences, being the concentrations at the outlet significantly lower than at the inlet, except in the case of SRP and NH4 + . O2 concentration at the inlet showed a large variability, being anoxic in several samplings. Vertical variations were observed at the outlet. On the surface, oxygen concentrations decreased from 4–5 mg l−1 until May 2003,
Outlet bottom Mean 16.9 1.8 0.5 7.5 296 53 538 4.5 0.27 4.2 6.5 0.67 3.0 6.1 <0.05 0.25 45 10
Range 8.0–27 1.2–2.9 0–7.5 7.0–8.3 195–595 36.1–77.2 158–950 1.0–17 0.02–1.3 0.13–18 2.0–21 0.11–3.2 1–3.5 3.5–8.1 <0.05 0.03–1.1 11–64 7–30
Mean removal (%)
37 65 44 70 60 −49 53 95 86 67 – −19 78 77
to 0–2 until August, and remained anoxic (<1 mg l−1 ) for the remainder of the study. The lower depths of the outlet were anoxic in most samplings. Oxygen concentrations induced differences between surface and bottom in most other parameters. Since the water released from the wetland was taken close to the bottom, removal efficiencies shown in Table 1 corresponded to the bottom samples. The design of the outlet facility is therefore important. Outcoming water quality would be improved just by taking it from the surface. BOD and COD were reduced by 77% and 78% at the outlet, suggesting a large decomposition and metabolism of incoming organic matter. Nitrate, nitrite and sulphate were reduced by 70%, 60% and 44%. On the contrary, ammonium concentration nearly doubled. Organic matter mineralization represents an important source of ammonium, which is not nitrified because low oxygen concentration limited nitrification. Due to nitrate in the incoming water was much greater than ammonium, the overall inorganic N balance showed a net reduction of 53% of the incoming inorganic N. Attained macrophyte biomass and tissue N concentrations suggest the biomass N pool represented <10% of the N removed from the incoming wastewater. Given the observed oxygen depletion it seems likely that denitrification is the major removal process. Several different studies have consistently shown denitrification to be a major pathway in wetlands. D’Angelo and Reddy (1993) determined that most of the 15 N-nitrate (roughly 90%) applied to sediment-water cores was lost by denitrification. Reddy et al. (1989) measured large denitrification rates in the rhizosphere of emergent macrophytes of deltaic wetlands. Matheson et al. (2002) performed 15 N balances in wetland microcosms estimating that denitrification accounted for 61% of the nitrate load, 25% was retained in the soil while only 14% was assimilated by the vegetation. Minzoni et al. (1988) measured large N losses through denitrification in enclosures installed on rice fields, and Golterman et al. (1988) confirmed the results by mass balances performed in experimental plots. Organic matter mineralization increased CO2 concentration in water, which, in turn, decreased water pH from 7.2–10.2
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Fig. 2 – pH, O2 , calcium, and SRP concentrations in the large-scale wetland during the study period.
at the inlet to 7.0–8.3 at the outlet. Mean calcium concentrations decreased 65% and alkalinity 37%. Removal percentages of calcium and alkalinity were greater when pH of the incoming water was higher (9.2–10.2). On the contrary, concentrations were slightly greater at the outlet when the pH of the incoming water was lower (7.2–7.6), suggesting that calcium carbonate precipitation within the wetland represents an important pathway governed by the incoming water pH. SRP concentrations at the inlet had greater standard deviations (Table 1). SRP concentrations at the inlet were lower when water pH was higher, above 9. At the outlet large differences between surface and bottom SRP concentrations were often observed, being greater at the bottom. Since the water released from the wetland was taken from the bottom, SRP concentration of the outcoming water had an overall mean concentration 19% greater than at the inlet. Mineralization of incoming organic matter seemed to be the main contribution to the observed increased concentration at the outlet. The low SRP concentrations at the inlet coincidentally with the samplings of high pH, calcium and carbonate might be caused by phosphorous coprecipitation with calcium carbonate (Fig. 2). As pH decreases, SRP sorption to carbonates decreases while adsorption to iron increases (Golterman, 1995). However, oxygen depletion prevented adsorption to iron, causing the observed large SRP concentrations in the outlet waters (Fig. 2).
The wetland showed high Fe, Cr and Ni retention (Table 1). The overall mean throughout the study period was 95%, 86% and 67% retention for Fe, Cr and Ni, respectively. The removal percentages of each metal remained almost constant during the experimental period, in consequence the highest the incoming concentrations, the largest the removed metal amounts. Zn concentration was below 50 g l−1 (detection limit of the analytical method) both in inlet and outlet throughout the study period. Simultaneous sulphate and Fe removal and oxygen depletion in the water column suggest insoluble FeS formation. Because of the high sulphate concentration in the incoming wastewater, most of the organic matter mineralization took place at the expense of biological sulphate reduction as observed in coastal marine sediments where sulphate reduction is responsible for 25–79% of the total organic matter mineralization. Hydrogen sulphide, liberated by sulphate reduction subsequently reacts with iron to form iron sulphide minerals (Giblin, 1988). Table 2 shows TP and metal concentrations in the bottom sediment. TP showed spatial and temporal variations. After the first year of operation, TP showed a significant increase at the inlet while at the outlet TP concentration was not significantly different comparing the beginning and the end of the study period. Greater sediment concentrations were consistent with the hypothesized SRP coprecipitation with carbonates at the inlet. Cr, Ni and Zn concentration in the bottom
e c o l o g i c a l e n g i n e e r i n g 2 6 ( 2 0 0 6 ) 341–347
Table 2 – Mean metal and TP concentration in the bottom sediment of the large-scale constructed wetland (units are g g−1 ) Initial
Final Inlet
Cr Ni Zn TP
30 19 60 343
± ± ± ±
5 4 7 12
25 17 53 426
± ± ± ±
Outlet
3 2 5 15
21 16 59 361
± ± ± ±
5 4 5 17
sediment did not increase significantly throughout the study period. Several macrophytes were transplanted into the wetland. E. crassipes (water hyacinth) became dominant and covered about 80% of the water surface until January 2004, when the wetland water was emptied out for 5 days for maintenance activities. Although some water remained at the bottom and the plants were anchored in the mud. Afterwards plant cover of water hyacinth decreased to 49% while plant height and chlorophyll concentration significantly decreased and became significantly smaller than in a nearby undisturbed floodplain wetland. After the maintenance activities, P. elephantipes (elephant panicgrass) grew among the water hyacinth plants and P. cordata (pickerelweed) decreased in cover until its total disappearance of the wetland. T. domingensis (cattail) increased progressively from 4 to 14% on the wetland surface and P. elephantipes remained covered about 2–4% along the study period. T. domingensis shoot density was lower while shoot height was higher than in the undisturbed floodplain wetland along most of 2003. However, shoot density and height did not show significant differences from the floodplain wetland in the late 2003–2004 growing period (December 2003–March 2004). Metal content in plant tissue increased significantly in E. crassipes and T. domingensis roots, but not in shoots (Table 3). Cr, Ni and Zn concentration in cattail roots were significantly greater at the end of the first year compared to the initial concentration. Cr and Ni concentrations in water hyacinth roots increased significantly, while Zn concentrations did not increase significantly by the end of the first year. Plant biomass attained 0.7–1.2 kg dw m−2 for water hyacinth and 1.9–4.0 kg dw m−2 for
Table 3 – Metals and TP in plant tissue at the inlet of the large-scale constructed wetland after 1 year operation (units are g g−1 ) Initial Roots
Final Leaves
Roots
Leaves
E. crassipes (water hyacinth) Cr 16 Ni 28 Zn 43 TP 1920
12 9 35 4190
78 42 24 1560
9 21 15 3020
T. domingensis (cattail) Cr 20 Ni 24 Zn 60 TP 1230
13 5 45 2980
139 198 129 1560
12 2 18 1980
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cattail. To ensure long-term removal, periodic harvests are being performed twice a year. Metal concentrations in the harvested plants are below the admissible limits for the national regulation for hazardous waste and fertilizers. Therefore these residues are being used as compost for ornamental plants cultivated in a greenhouse which was constructed in the same factory. Metals in ornamental plant tissues will be monitored to ensure maximum safety conditions. The large-scale wetland showed retention efficiency for metals similar to that of the small-scale prototype studied earlier (Table 4). The incoming wastewater was different; the influent of the small-scale wetland presented greater pH, conductivity, nutrient and metal concentrations than the larger wetland (Table 4). Three floating and eight emergent macrophytes were transplanted to the small-scale wetland. After some initial growth, the small-scale wetland became a monospecific stand of cattail with a biomass similar to undisturbed environments (Maine et al., 2005). Metal concentration in both, incoming wastewater and plant tissues, were lower than the metal thresholds reported in literature (Gibson and Pollard, 1988; Delgado et al., 1993; Sen and Bhattacharyya, 1994; Selvaphathy et al., 1997; Cardwell et al., 2002; Fakayode and Onianwa, 2002; Ingole and Bhole, 2003; Manios et al., 2003; Soltan and Rashed, 2003, Maine et al., 2004). Hadad et al. (in press) reported pH toxicity thresholds of 10, 11 and 11 for E. crassipes, P. stratiotes and Salvinia herzogii, respectively, and conductivity toxicity thresholds of 4 and 8 mS cm−1 for E. crassipes and P. stratiotes, respectively. In most of the samplings in the small-scale wetland conductivity and pH were higher than the tolerance thresholds (Maine et al., 2005). In consequence, conductivity and pH were probably the main cause of the progressive disappearance of most transplanted species. As regards emergent species, Klomjek and Nitisoravut (2005) studied the response given by eight species used in a constructed wetland, similar to the one studied at a pilot scale which received saline wastewaters (14–16 mS cm−1 ). T. angustifolia, among others, proved to be tolerant and showed a satisfactory growth. These conductivity values are higher than those recorded for the studied wetland, being this the reason why T. domingensis adapted to the conditions of the system, presenting a positive growth rate. It seems therefore not surprising that floating macrophytes disappeared comparatively earlier than emergent ones. In consequence, the floating macrophytes could not develop within the pH and conductivity prevailing in the incoming wastewater of the small-scale wetland even in the absence of metals. Increased water depth and decreased pH, conductivity and metal concentration in the large-scale wetland resulted in the dominance of water hyacinth. Oxygen concentrations in the large-scale constructed wetland were consistently lower than those of the small-scale wetland, in spite of lower COD and BOD loads in the former. Extensive development of water hyacinth causes oxygen depletion irrespective of external loadings as observed in many natural undisturbed floodplain environments of the Parana´ River (Pedrozo et al., 1992). Oxygen concentration seems to be the cause of different retention mechanisms between the two wetlands: while Cr, Ni, and Zn were mainly stored in the bottom sediment in the small-scale wetland, in the large-scale facility metals were retained in the vegetation. Guo et al. (1997a,b) studied metal speciation at differ-
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Table 4 – Water composition at the inlet and outlet and removal efficiency of the small-scale wetland (Maine et al., 2005) Inlet Mean ◦
Temperature ( C) Conductivity (mS cm−1 ) O2 (mg l−1 ) pH Alkalinity as CaCO3 (mg l−1 ) Ca (mg l−1 ) SO4 2− (mg l−1 ) N–NO3 − (mg l−1 ) N–NO2 − (mg l−1 ) N–NH4 + (mg l−1 ) Inorganic N (mg l−1 ) Fe (mg l−1 ) Cr (g l−1 ) Ni (g l−1 ) Zn (g l−1 ) SRP (g l−1 ) COD (mg l−1 ) BOD (mg l−1 )
18.4 5.13 6.18 10 303 171 1444 20 2.25 2.2 24.8 9.1 127 181 60 550 276 136
Outlet Range
Mean
Range
11–26 3.3–8.5 0–10.6 8.0–12.5 179–539 43–292 298–2322 5–86 0–12 0.3–4.9 10–95 0.05–32 5–589 3–750 40–210 30–2600 57–583 17–400
18.4 3.1 3.15 7.9 344 97 929 1.5 0.15 2.2 3.91 0.21 13 51 40 141 40 15
11–26 0.7–5.4 0–5.8 7.0–9.0 166–665 22–238 204–1690 0.2–3.9 0.001–0.89 0.006–9.5 0.31–14.4 0.05–0.7 1–105 4–190 10–69 10–572 22–85 2–38
ent redox potentials in sediment-water suspensions. At high redox potentials (430 mV) Cr, Ni, and Zn are adsorbed onto Fe and Mn colloids. With the oxygen concentrations that prevailed in the small-scale wetland coprecipitation with iron determined the metal retention within the bottom sediment. On the contrary, almost permanent oxygen depletion in the large-scale wetland prevented precipitation to the wetland bottom. Macrophyte roots release oxygen to the rhizosphere (Reddy et al., 1989) and produce the precipitation of Fe to form the so-called “iron-plaque” (Otte et al., 1995). Metal binding affinity to iron oxyhydroxides causes metal accumulation into the iron-plaque (Otte et al., 1995). Thus, metals might either be directly sorbed by the macrophytes from the solution or co-precipitate with Fe onto the macrophyte roots. Oxygen depletion also seemed to be the cause of the increased SRP and ammonium at the outlet while the small-scale wetland efficiently removed SRP.
4.
% Removal
Conclusions
• Both, the preliminary small-scale and the subsequent largescale facility efficiently removed metals from the effluent of a metallurgic plant. The small-scale and the larger wetlands reduced Cr, Ni, Fe concentrations by 81%, 66%, 82%, and 86%, 67%, 95%, respectively. However, the small-scale was equally efficient in removing SRP while the large-scale was not. • Conductivity and pH of the influent were much greater in the small-scale wetland causing the disappearance of the floating macrophytes, while the large-scale was mainly covered by water hyacinth. Replacing the present dominance of water hyacinth by emergent macrophytes would help to increase dissolved oxygen concentrations. A shift from the removal mechanism based on macrophyte absorption attained in the large scale wetland to a bottom coprecipitation attained in the small-scale wetland might be induced by manipulating macrophyte dominance.
– – – – – 41 35 88 85 13 84 82 81 66 59 75 86 89
• Decreasing water level will favour the dominance of emergent macrophytes as evidenced by the changes produced when the wetland was emptied for maintenance. Furthermore, reducing water depth would contribute to increasing oxygen concentration.
Acknowledgements The authors thank Consejo Nacional de Investigaciones ´ Cient´ıficas y Tecnicas (CONICET) from Argentina and Universidad Nacional del Litoral, CAI + D Project for providing funds for this work.
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sewage sludge compost and watered with metaliferus water. Ecol. Eng. 20, 65–74. Matheson, F., Nguyen, M., Cooper, A., Burt, T., Bull, E., 2002. Fate of 15 N-nitrate in unplanted, planted and harvested riparian wetland soil microcosms. Ecol. Eng. 19, 249–264. Minzoni, F., Bonetto, C., Golterman, H., 1988. The nitrogen cycle in shallow water sediment systems of rice fields. Part I. The denitrification process. Hydrobiologia 159, 189–202. Murphy, J., Riley, J., 1962. A modified single solution method for determination of phosphate in natural waters. Anal. Chim. Acta 27, 31–36. Otte, M., Kearns, C., Doyle, M., 1995. Accumulation of arsenic and zinc in the rhizosphere of wetland plants. Bull. Environ. Contam. Toxicol. 55, 154–161. Pedrozo, F., Diaz, M., Bonetto, C., 1992. Nitrogen and phosphorus in the Parana´ River floodplain waterbodies. Arch. Hydrobiol. 2, 171–181. Reddy, K., Patrick, W., Lindau, C., 1989. Nitrification-denitrification at the plant root–sediment interface in wetlands. Limnol. Oceanogr. 34 (6), 1004–1013. Selvaphathy, P., Juliet Jesline, J., Prebha, S., 1997. Heavy metal removal from wastewater by water lettuce. Indian J. Environ. Prot. 18, 1–6. Sen, A.K., Bhattacharyya, M., 1994. Studies of uptake and toxic efects of Ni(II) on Salvinia natans. Water Air Soil Poll. 78, 141–152. Soltan, M., Rashed, M., 2003. Laboratory study on the survival of water hyacinth under several conditions of heavy metal concentrations. Adv. Environ. Res. 7, 321–334. Vesk, P., Allaway, W., 1997. Spatial variation of copper and lead concentrations of water hyacinth plants in a wetland receiving urban run-off. Aquat. Bot. 59, 33–44. Vollenweider, R.A., 1974. A Manual on Methods for Measuring. Primary Production in Aquatic Environments. IBP Handbook No. 12. International Biological Programme, Blackwell Scientific Publications, Oxford, 225 pp.
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Nutrient and metal removal in a constructed wetland for wastewater treatment from a metallurgic industry a ˜ a , H. Hadad a , G. Sanchez ´ M.A. Maine a,∗ , N. Sune , C. Bonetto b a
Qu´ımica Anal´ıtica, Facultad de Ingenier´ıa Qu´ımica, Universidad Nacional del Litoral, Santiago del Estero 2829 (3000), Santa Fe, Argentina b Instituto de Limnolog´ıa “Dr. Ringuelet”, Av. Calchaqui km 23.5 (1888), Florencio Varela, Buenos Aires, Argentina
a r t i c l e
i n f o
a b s t r a c t
Article history:
This contribution summarizes the nutrient and metal removal of a free water surface con-
Received 1 April 2005
structed wetland, compares it with the previous small-scale prototype and discusses the
Received in revised form 9
observed differences. Several locally available macrophyte species were transplanted into
December 2005
the wetland. Eichhornia crassipes (water hyacinth) showed a fast growth and it soon became
Accepted 28 December 2005
dominant, attaining 80% cover of the wetland surface. Typha domingensis (cattail) and Panicum elephantipes (elephant panicgrass) developed as accompanying species attaining 14 and 4% cover. The wetland removed 86% of Cr and 67% of Ni. Zn concentrations were below
Keywords:
50 g l−1 in most samplings. The FeS precipitation probably caused the high retention of
Metal
Fe (95%). The outcoming water was anoxic in most samplings. Phosphate and ammonium
Nutrient
were not retained within the wetland while 70% and 60% of the incoming nitrate and nitrite
Removal
were removed. Large denitrification losses are suggested. Cr, Ni and Zn were retained by the
Wastewater
macrophytes in the larger wetland and in sediment in the small-scale one. Differences in
Wetland
the retention mechanism of the two wetlands are discussed. © 2006 Elsevier B.V. All rights reserved.
1.
Introduction
The construction of artificial wetlands for wastewater treatment is now a widely accepted and increasingly common treatment alternative. Constructed wetlands were initially utilized for nutrient removal in residential and municipal sewage, storm water and agricultural runoff displaying a wide range of removal efficiencies. The accelerating industrialization in developing countries with an enormous consumption of metals constitutes an environmental contamination hazard. The application of wetlands for industrial wastewater treatment is a promising alternative. However, current experience in Argentina remains largely unreported. Conditions are favourable since there is a large availability of marginal land around most cities with a low population density. The central and northern areas of the country have mild winters, allow-
∗
Corresponding author. Tel.: +54 342 4571164; fax: +54 342 4571162. E-mail address: amaine@fiqus.unl.edu.ar (M.A. Maine). 0925-8574/$ – see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2005.12.004
ing extended growing periods for plants. Macrophytes are the main biological component of wetlands. They not only assimilate pollutants directly into their tissues, but they also act as catalysts for purification reactions by increasing the environment diversity in the root zone and promoting a variety of chemical and biochemical reactions which enhance purification (Jenssen et al., 1993). Water hyacinth (Eichhornia crassipes) is one of the most commonly used plants in constructed wetlands because of its fast growth rate and large uptake of nutrients and contaminants. It attains dense stands in the floodplain wetlands of the Middle Parana´ River close to the study site. Bahco S.A., a metallurgic factory, constructed a small-scale experimental free water surface wetland to assess the feasibility of treating wastewater from the Santo Tome´ (Argentina) tool manufacturing plant. The wetland removed 81%, 66% and
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59% of the incoming Cr, Ni and Zn, respectively, and 84% and 75% of the inorganic N and soluble reactive P (SRP) from the incoming wastewater (Maine et al., 2005). Because of the high removal efficiency attained in the small-scale experimental wetland, a large-scale facility was constructed for wastewater treatment of the whole factory. The present contribution assesses the removal efficiency of the larger wetland, compares it with the previous small-scale prototype and discusses the observed differences.
2.
Materials and methods
2.1. Design of the small-scale and the larger constructed wetlands The wetlands were constructed at Bahco Argentina metal´ Argentina (S31◦ 40 01.9 ; lurgic plant, located in Santo Tome, W60◦ 47 06.9 ). The small-scale experimental wetland was 6 m length, 3 m wide and 0.4 m deep (Fig. 1A). A polyethylene impermeable film was placed at the bottom and a soil layer of 30 cm was added. The influent entered the wetland through a PVC tube (diameter: 63 mm) with a perpendicular drip dispersion tube with aligned holes to produce a laminar flow. The inflow discharge was 1000 l d−1 and the approximate hydraulic residence time was 7 days. Three floating (Pistia stratiotes, E. crassipes and Salvinia rotundifolia) and eight emergent (Cyperus alternifolius, P. elephantipes, Thalia geniculata, Polygonum punctatum, Pontederia cordata, Pontederia rotundifolia, Typha domingensis and Aechmea distichantia) macrophytes were transplanted (Maine et al., 2005). The larger free water surface wetland was 50 m length by 40 m wide and 0.5–0.8 m deep, with a central baffle (Fig. 1B). The baffle doubled the flowpath and resulted in a 5:1 length–width ratio. The wetland received wastewater through a PVC pipe provided with a perpendicular distribution pipe with holes at regular distances in order to allow uniform distribution of flow. Wastewater discharge was approximately 100 m3 d−1 and the hydraulic residence time ranged from 7
Fig. 1 – Scheme of the (A) small-scale wetland and (B) large-scale wetland.
to 12 days. The wetland was rendered impermeable by means of a bentonite layer covered with a 1 m-layer of the excavated soil. Several locally available macrophytes were transplanted into the wetland, being E. crassipes, T. domingensis and P. cordata those of the greatest cover. After crossing the wetland, the effluent followed an excavated channel to a 1.5 ha pond. Both wastewater from the industrial processes and sewage from the factory were treated together. It was expected that high nutrient concentrations could increase the toxicity tolerance of the macrophytes (Manios et al., 2003). Effluents reached the wetland after a primary treatment. During the first 3 months of operation only sewage entered the wetland. Later, industrial waste plus sewage were treated.
2.2.
Sampling
Twenty-four water samplings were performed at the inlet and outlet of the larger free water surface wetland from November 2002 through February 2004. Samples were taken monthly until March 2003, and every 2 weeks afterwards. Sediment total P (TP), Ni, Cr and Zn concentrations were determined twice in a month at the inlet and outlet of the wetland. Sediment samples were collected using a 4 cm-diameter PVC corer. All the samples were transported to the laboratory at 4 ◦ C. In order to estimate the biomass, emergent and floating macrophyte stands were sampled with a 0.50 m × 0.50 m sampler following Vesk and Allaway (1997). Four replicates were taken. Macrophytes were then harvested and sorted by species at the laboratory, washed, separated between aboveground (shoots and leaves) and belowground (roots and rhizomes) parts, and dried at 105 ◦ C until constant weight was reached (APHA, 1998). Plant cover was estimated measuring the area occupied by each stand within the wetland.
2.3.
Chemical analysis
Conductivity was measured with a 33 model YSI conductimeter, O2 with a Horiba OM-14 portable meter and pH with an Orion pH-meter. Water samples were filtered through Millipore membrane filters (0.45 m) for P and N determinations. Chemical analysis was performed following APHA (1998); NO2 − was determined by coupling diazotation followed by a colorimetric technique, NH4 + and NO3 − by potentiometry (Orion ion selective electrodes, sensitivity: 0.01 mg l−1 of N, reproducibility: ±2%). Soluble reactive phosphate (SRP) was determined by the colorimetric molybdenum blue method (Murphy and Riley, 1962). Ca2+ and Mg2+ were determined by EDTA titration. Na+ and K+ were determined by flame emission photometry. Alkalinity (CO3 2− and HCO3 − ) was measured by HCl titration. Cl− was determined by the argentometric method. SO4 2− was assessed by turbidimetry. COD was determined by the open reflux method and BOD by the 5-day BOD test (APHA, 1998). Fe, Cr, Ni and Zn concentrations were determined in water samples by atomic absorption spectrometry (by flame or electrothermal atomization, according to the sample concentration, Perkin-Elmer 5000), following APHA (1998). Total phosphorous (TP) in sediment was determined after an acid digestion with an HClO4 /HNO3 /HCl (7:5:2) mixture followed by SRP determination in the digested samples (Murphy
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e c o l o g i c a l e n g i n e e r i n g 2 6 ( 2 0 0 6 ) 341–347
Table 1 – Measured variables in the constructed wetland during the study period Inlet
Parameter
◦
Temperature ( C) Conductivity (mS cm−1 ) O2 (mg l−1 ) pH Alkalinity (mg l−1 CaCO3 ) Ca (mg l−1 ) SO4 2− (mg l−1 ) N–NO3 − (mg l−1 ) N–NO2 − (mg l−1 ) N–NH4 + (mg l−1 ) Inorganic N (mg l−1 ) Fe (mg l−1 ) Cr (g l−1 ) Ni (g l−1 ) Zn (g l−1 ) SRP (mg l−1 ) COD (mg l−1 ) BOD (mg l−1 )
Outlet surface
Mean
Range
Mean
Range
16.9 2.9 2.4 8.7 463 156 957 16 0.79 1.9 19 13.7 22 17 <0.05 0.20 204 41
8.0–27 0.4–8.5 0–7.1 7.2–10.8 205–1187 27.0–651 98.1–2506 1.6–68 0.021–3.4 0.1–12 1.8–74 0.16–74 3.3–150 6.1–60 <0.05 0.003–0.51 22–430 7–89
16.9 1.3 1.5 7.2 237 43 395 3.0 0.15 1.6 5.0 0.38 3.6 9.0 <0.05 0.19 39 15
8.0–27 0.47–2.6 0–5.6 6.9–8.1 95–475 22.3–61.2 159–855 0.8–7.9 0.01–0.99 0.12–6.7 1.2–10 0.05–1.2 2.8–5.3 3.9–27 <0.05 0.02–0.43 15–103 5–32
and Riley, 1962). Cr, Ni and Zn sediment concentrations were determined in the digests by atomic absorption spectrometry (Perkin-Elmer 5000). Above and belowground macrophyte tissues were ground for P and metal determination. TP, Cr, Ni and Zn in plant tissues were determined in the same way as in sediment samples. Chlorophyll was extracted with acetone for 48 h under dark and cold conditions (3–5 ◦ C). The transmittance percentage of the extracts was read to 645 and 665 nm using a UV–vis spectrophotometer following Vollenweider (1974) in order to calculate chlorophyll a concentration. A nearby undisturbed floodplain wetland of the Parana´ river was also sampled to compare plant height, chlorophyll and nutrient contents in plants with those of the wastewater wetland.
2.4.
Statistical analysis
Statistical significance between inlet and outlet water concentrations was assessed using a paired-sample comparison test (p < 0.05). Differences in the concentrations of TP and metals between inlet and outlet sediment and between initial and final concentrations were determined by a mean comparison test (p < 0.05). Calculations were carried out with the Statgraphics Plus 3.0 program.
3.
Results and discussion
Table 1 summarizes the variables measured at the inlet and outlet of the wetland and the estimated removal efficiency. Comparing the concentration of variables measured at the inlet and the outlet in the different sampling, there are significantly statistical differences, being the concentrations at the outlet significantly lower than at the inlet, except in the case of SRP and NH4 + . O2 concentration at the inlet showed a large variability, being anoxic in several samplings. Vertical variations were observed at the outlet. On the surface, oxygen concentrations decreased from 4–5 mg l−1 until May 2003,
Outlet bottom Mean 16.9 1.8 0.5 7.5 296 53 538 4.5 0.27 4.2 6.5 0.67 3.0 6.1 <0.05 0.25 45 10
Range 8.0–27 1.2–2.9 0–7.5 7.0–8.3 195–595 36.1–77.2 158–950 1.0–17 0.02–1.3 0.13–18 2.0–21 0.11–3.2 1–3.5 3.5–8.1 <0.05 0.03–1.1 11–64 7–30
Mean removal (%)
37 65 44 70 60 −49 53 95 86 67 – −19 78 77
to 0–2 until August, and remained anoxic (<1 mg l−1 ) for the remainder of the study. The lower depths of the outlet were anoxic in most samplings. Oxygen concentrations induced differences between surface and bottom in most other parameters. Since the water released from the wetland was taken close to the bottom, removal efficiencies shown in Table 1 corresponded to the bottom samples. The design of the outlet facility is therefore important. Outcoming water quality would be improved just by taking it from the surface. BOD and COD were reduced by 77% and 78% at the outlet, suggesting a large decomposition and metabolism of incoming organic matter. Nitrate, nitrite and sulphate were reduced by 70%, 60% and 44%. On the contrary, ammonium concentration nearly doubled. Organic matter mineralization represents an important source of ammonium, which is not nitrified because low oxygen concentration limited nitrification. Due to nitrate in the incoming water was much greater than ammonium, the overall inorganic N balance showed a net reduction of 53% of the incoming inorganic N. Attained macrophyte biomass and tissue N concentrations suggest the biomass N pool represented <10% of the N removed from the incoming wastewater. Given the observed oxygen depletion it seems likely that denitrification is the major removal process. Several different studies have consistently shown denitrification to be a major pathway in wetlands. D’Angelo and Reddy (1993) determined that most of the 15 N-nitrate (roughly 90%) applied to sediment-water cores was lost by denitrification. Reddy et al. (1989) measured large denitrification rates in the rhizosphere of emergent macrophytes of deltaic wetlands. Matheson et al. (2002) performed 15 N balances in wetland microcosms estimating that denitrification accounted for 61% of the nitrate load, 25% was retained in the soil while only 14% was assimilated by the vegetation. Minzoni et al. (1988) measured large N losses through denitrification in enclosures installed on rice fields, and Golterman et al. (1988) confirmed the results by mass balances performed in experimental plots. Organic matter mineralization increased CO2 concentration in water, which, in turn, decreased water pH from 7.2–10.2
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Fig. 2 – pH, O2 , calcium, and SRP concentrations in the large-scale wetland during the study period.
at the inlet to 7.0–8.3 at the outlet. Mean calcium concentrations decreased 65% and alkalinity 37%. Removal percentages of calcium and alkalinity were greater when pH of the incoming water was higher (9.2–10.2). On the contrary, concentrations were slightly greater at the outlet when the pH of the incoming water was lower (7.2–7.6), suggesting that calcium carbonate precipitation within the wetland represents an important pathway governed by the incoming water pH. SRP concentrations at the inlet had greater standard deviations (Table 1). SRP concentrations at the inlet were lower when water pH was higher, above 9. At the outlet large differences between surface and bottom SRP concentrations were often observed, being greater at the bottom. Since the water released from the wetland was taken from the bottom, SRP concentration of the outcoming water had an overall mean concentration 19% greater than at the inlet. Mineralization of incoming organic matter seemed to be the main contribution to the observed increased concentration at the outlet. The low SRP concentrations at the inlet coincidentally with the samplings of high pH, calcium and carbonate might be caused by phosphorous coprecipitation with calcium carbonate (Fig. 2). As pH decreases, SRP sorption to carbonates decreases while adsorption to iron increases (Golterman, 1995). However, oxygen depletion prevented adsorption to iron, causing the observed large SRP concentrations in the outlet waters (Fig. 2).
The wetland showed high Fe, Cr and Ni retention (Table 1). The overall mean throughout the study period was 95%, 86% and 67% retention for Fe, Cr and Ni, respectively. The removal percentages of each metal remained almost constant during the experimental period, in consequence the highest the incoming concentrations, the largest the removed metal amounts. Zn concentration was below 50 g l−1 (detection limit of the analytical method) both in inlet and outlet throughout the study period. Simultaneous sulphate and Fe removal and oxygen depletion in the water column suggest insoluble FeS formation. Because of the high sulphate concentration in the incoming wastewater, most of the organic matter mineralization took place at the expense of biological sulphate reduction as observed in coastal marine sediments where sulphate reduction is responsible for 25–79% of the total organic matter mineralization. Hydrogen sulphide, liberated by sulphate reduction subsequently reacts with iron to form iron sulphide minerals (Giblin, 1988). Table 2 shows TP and metal concentrations in the bottom sediment. TP showed spatial and temporal variations. After the first year of operation, TP showed a significant increase at the inlet while at the outlet TP concentration was not significantly different comparing the beginning and the end of the study period. Greater sediment concentrations were consistent with the hypothesized SRP coprecipitation with carbonates at the inlet. Cr, Ni and Zn concentration in the bottom
e c o l o g i c a l e n g i n e e r i n g 2 6 ( 2 0 0 6 ) 341–347
Table 2 – Mean metal and TP concentration in the bottom sediment of the large-scale constructed wetland (units are g g−1 ) Initial
Final Inlet
Cr Ni Zn TP
30 19 60 343
± ± ± ±
5 4 7 12
25 17 53 426
± ± ± ±
Outlet
3 2 5 15
21 16 59 361
± ± ± ±
5 4 5 17
sediment did not increase significantly throughout the study period. Several macrophytes were transplanted into the wetland. E. crassipes (water hyacinth) became dominant and covered about 80% of the water surface until January 2004, when the wetland water was emptied out for 5 days for maintenance activities. Although some water remained at the bottom and the plants were anchored in the mud. Afterwards plant cover of water hyacinth decreased to 49% while plant height and chlorophyll concentration significantly decreased and became significantly smaller than in a nearby undisturbed floodplain wetland. After the maintenance activities, P. elephantipes (elephant panicgrass) grew among the water hyacinth plants and P. cordata (pickerelweed) decreased in cover until its total disappearance of the wetland. T. domingensis (cattail) increased progressively from 4 to 14% on the wetland surface and P. elephantipes remained covered about 2–4% along the study period. T. domingensis shoot density was lower while shoot height was higher than in the undisturbed floodplain wetland along most of 2003. However, shoot density and height did not show significant differences from the floodplain wetland in the late 2003–2004 growing period (December 2003–March 2004). Metal content in plant tissue increased significantly in E. crassipes and T. domingensis roots, but not in shoots (Table 3). Cr, Ni and Zn concentration in cattail roots were significantly greater at the end of the first year compared to the initial concentration. Cr and Ni concentrations in water hyacinth roots increased significantly, while Zn concentrations did not increase significantly by the end of the first year. Plant biomass attained 0.7–1.2 kg dw m−2 for water hyacinth and 1.9–4.0 kg dw m−2 for
Table 3 – Metals and TP in plant tissue at the inlet of the large-scale constructed wetland after 1 year operation (units are g g−1 ) Initial Roots
Final Leaves
Roots
Leaves
E. crassipes (water hyacinth) Cr 16 Ni 28 Zn 43 TP 1920
12 9 35 4190
78 42 24 1560
9 21 15 3020
T. domingensis (cattail) Cr 20 Ni 24 Zn 60 TP 1230
13 5 45 2980
139 198 129 1560
12 2 18 1980
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cattail. To ensure long-term removal, periodic harvests are being performed twice a year. Metal concentrations in the harvested plants are below the admissible limits for the national regulation for hazardous waste and fertilizers. Therefore these residues are being used as compost for ornamental plants cultivated in a greenhouse which was constructed in the same factory. Metals in ornamental plant tissues will be monitored to ensure maximum safety conditions. The large-scale wetland showed retention efficiency for metals similar to that of the small-scale prototype studied earlier (Table 4). The incoming wastewater was different; the influent of the small-scale wetland presented greater pH, conductivity, nutrient and metal concentrations than the larger wetland (Table 4). Three floating and eight emergent macrophytes were transplanted to the small-scale wetland. After some initial growth, the small-scale wetland became a monospecific stand of cattail with a biomass similar to undisturbed environments (Maine et al., 2005). Metal concentration in both, incoming wastewater and plant tissues, were lower than the metal thresholds reported in literature (Gibson and Pollard, 1988; Delgado et al., 1993; Sen and Bhattacharyya, 1994; Selvaphathy et al., 1997; Cardwell et al., 2002; Fakayode and Onianwa, 2002; Ingole and Bhole, 2003; Manios et al., 2003; Soltan and Rashed, 2003, Maine et al., 2004). Hadad et al. (in press) reported pH toxicity thresholds of 10, 11 and 11 for E. crassipes, P. stratiotes and Salvinia herzogii, respectively, and conductivity toxicity thresholds of 4 and 8 mS cm−1 for E. crassipes and P. stratiotes, respectively. In most of the samplings in the small-scale wetland conductivity and pH were higher than the tolerance thresholds (Maine et al., 2005). In consequence, conductivity and pH were probably the main cause of the progressive disappearance of most transplanted species. As regards emergent species, Klomjek and Nitisoravut (2005) studied the response given by eight species used in a constructed wetland, similar to the one studied at a pilot scale which received saline wastewaters (14–16 mS cm−1 ). T. angustifolia, among others, proved to be tolerant and showed a satisfactory growth. These conductivity values are higher than those recorded for the studied wetland, being this the reason why T. domingensis adapted to the conditions of the system, presenting a positive growth rate. It seems therefore not surprising that floating macrophytes disappeared comparatively earlier than emergent ones. In consequence, the floating macrophytes could not develop within the pH and conductivity prevailing in the incoming wastewater of the small-scale wetland even in the absence of metals. Increased water depth and decreased pH, conductivity and metal concentration in the large-scale wetland resulted in the dominance of water hyacinth. Oxygen concentrations in the large-scale constructed wetland were consistently lower than those of the small-scale wetland, in spite of lower COD and BOD loads in the former. Extensive development of water hyacinth causes oxygen depletion irrespective of external loadings as observed in many natural undisturbed floodplain environments of the Parana´ River (Pedrozo et al., 1992). Oxygen concentration seems to be the cause of different retention mechanisms between the two wetlands: while Cr, Ni, and Zn were mainly stored in the bottom sediment in the small-scale wetland, in the large-scale facility metals were retained in the vegetation. Guo et al. (1997a,b) studied metal speciation at differ-
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26
( 2 0 0 6 ) 341–347
Table 4 – Water composition at the inlet and outlet and removal efficiency of the small-scale wetland (Maine et al., 2005) Inlet Mean ◦
Temperature ( C) Conductivity (mS cm−1 ) O2 (mg l−1 ) pH Alkalinity as CaCO3 (mg l−1 ) Ca (mg l−1 ) SO4 2− (mg l−1 ) N–NO3 − (mg l−1 ) N–NO2 − (mg l−1 ) N–NH4 + (mg l−1 ) Inorganic N (mg l−1 ) Fe (mg l−1 ) Cr (g l−1 ) Ni (g l−1 ) Zn (g l−1 ) SRP (g l−1 ) COD (mg l−1 ) BOD (mg l−1 )
18.4 5.13 6.18 10 303 171 1444 20 2.25 2.2 24.8 9.1 127 181 60 550 276 136
Outlet Range
Mean
Range
11–26 3.3–8.5 0–10.6 8.0–12.5 179–539 43–292 298–2322 5–86 0–12 0.3–4.9 10–95 0.05–32 5–589 3–750 40–210 30–2600 57–583 17–400
18.4 3.1 3.15 7.9 344 97 929 1.5 0.15 2.2 3.91 0.21 13 51 40 141 40 15
11–26 0.7–5.4 0–5.8 7.0–9.0 166–665 22–238 204–1690 0.2–3.9 0.001–0.89 0.006–9.5 0.31–14.4 0.05–0.7 1–105 4–190 10–69 10–572 22–85 2–38
ent redox potentials in sediment-water suspensions. At high redox potentials (430 mV) Cr, Ni, and Zn are adsorbed onto Fe and Mn colloids. With the oxygen concentrations that prevailed in the small-scale wetland coprecipitation with iron determined the metal retention within the bottom sediment. On the contrary, almost permanent oxygen depletion in the large-scale wetland prevented precipitation to the wetland bottom. Macrophyte roots release oxygen to the rhizosphere (Reddy et al., 1989) and produce the precipitation of Fe to form the so-called “iron-plaque” (Otte et al., 1995). Metal binding affinity to iron oxyhydroxides causes metal accumulation into the iron-plaque (Otte et al., 1995). Thus, metals might either be directly sorbed by the macrophytes from the solution or co-precipitate with Fe onto the macrophyte roots. Oxygen depletion also seemed to be the cause of the increased SRP and ammonium at the outlet while the small-scale wetland efficiently removed SRP.
4.
% Removal
Conclusions
• Both, the preliminary small-scale and the subsequent largescale facility efficiently removed metals from the effluent of a metallurgic plant. The small-scale and the larger wetlands reduced Cr, Ni, Fe concentrations by 81%, 66%, 82%, and 86%, 67%, 95%, respectively. However, the small-scale was equally efficient in removing SRP while the large-scale was not. • Conductivity and pH of the influent were much greater in the small-scale wetland causing the disappearance of the floating macrophytes, while the large-scale was mainly covered by water hyacinth. Replacing the present dominance of water hyacinth by emergent macrophytes would help to increase dissolved oxygen concentrations. A shift from the removal mechanism based on macrophyte absorption attained in the large scale wetland to a bottom coprecipitation attained in the small-scale wetland might be induced by manipulating macrophyte dominance.
– – – – – 41 35 88 85 13 84 82 81 66 59 75 86 89
• Decreasing water level will favour the dominance of emergent macrophytes as evidenced by the changes produced when the wetland was emptied for maintenance. Furthermore, reducing water depth would contribute to increasing oxygen concentration.
Acknowledgements The authors thank Consejo Nacional de Investigaciones ´ Cient´ıficas y Tecnicas (CONICET) from Argentina and Universidad Nacional del Litoral, CAI + D Project for providing funds for this work.
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