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Redox potential dynamics in a horizontal subsurface flow constructed wetland for wastewater treatment: Diel, seasonal and spatial fluctuations
e c o l o g i c a l e n g i n e e r i n g 3 4 ( 2 0 0 8 ) 223–232
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Redox potential dynamics in a horizontal subsurface flow constructed wetland for wastewater treatment: Diel, seasonal and spatial fluctuations a,c ˇ zková ˇ a,b,∗ , Tomásˇ Picek b , Hana Cí ˇ Jiˇrí Dusek a b c
ˇ Czech Republic Wetland Ecology, Institute of Systems Biology and Ecology, Tˇrebon, ˇ Department of Ecosystem Biology, University of South Bohemia, Faculty of Science, Ceské Budˇejovice, Czech Republic ˇ Department of Biology, University of South Bohemia, Faculty of Agriculture, Ceské Budˇejovice, Czech Republic
a r t i c l e
i n f o
a b s t r a c t
Article history:
Redox potential was measured in a treatment bed of a constructed wetland (CW) with sub-
Received 20 May 2008
surface horizontal flow used for municipal wastewater treatment. CW was situated in the
Received in revised form
Czech Republic, designed for 150 person equivalents; it was planted with Phragmites aus-
4 August 2008
tralis and put into operation in 2001. The system was very efficient in removing organic
Accepted 5 August 2008
pollution (BOD5 82% and COD 74.0%), nitrogen (62.6%), phosphorus (75.4%) and suspended solids (52%). Redox potential (EH ) was measured in two depths (0.2 and 0.5 m) of the treatment bed at regular distances from inflow using platinum electrodes, continuously, every
Keywords:
15 min during a more-than-2-year period (27 months) from 2002 to 2004. EH ranged from
Redox potential
−400 to +800 mV. The fluctuations were the most intense in the upper layer (0.2 m) at the
Redox processes
furthest distance from inflow (13 m) and EH changed within very short time periods (hours)
Phragmites australis
by several hundred millivolts. Among the factors affecting EH concentrations of pollutants,
Wastewater treatment
pore water temperature and water flow rate were determined.
Constructed wetlands Continuous measurement
Vegetated treatment beds of subsurface horizontal flow constructed wetlands filled with mineral substrate can be extremely dynamic systems, in which oxidation–reduction status changes within short time from anaerobic to aerobic and vice versa. Anaerobic conditions do not prevail necessarily in the upper layer of treatment bed, in which plant roots are present. © 2008 Elsevier B.V. All rights reserved.
1.
Introduction
Constructed wetlands (CWs) are becoming increasingly popular worldwide for removing organic matter, nutrients, trace elements, pathogens, or other pollutants from wastewater and/or runoff. CWs used for wastewater treatment provide a relatively simple and inexpensive solution for the treatment of wastewaters from small communities and industries, as well as storm sewage, and agricultural runoff (Kadlec and Knight, 1996). These treatment systems are engineered systems that
∗
Corresponding author. Tel.: +420 384706182. E-mail address: [email protected] (J. Duˇsek). URL: http://www.usbe.cas.cz/index.php?node=466 (J. Duˇsek).
0925-8574/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2008.08.008
have been designed and constructed to utilize natural processes (Mitsch and Gosselink, 1986; Brix, 1989). The natural processes involve wetland vegetation, soils, and their associated microbial assemblages to assist in treating wastewater. Most existing CWs in the Czech Republic are systems with subsurface horizontal wastewater flow. These systems are designed to remove organics (BOD5 and COD) and suspended solids as those parameters are usually limiting in the effluent. The systems treat municipal or domestic sewage often combined with storm water runoff. The treatment beds of these
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systems are usually planted with common reed (Phragmites australis Trin. Ex Steudel), reed canary grass (Phalaris arundinacea L.) or with both plant species. The growth of common reed and reed canary grass in the CWs was described by Duˇsek ˇ (2001), Duˇsek et al. (2001), and Vymazal (1997), Duˇsek and Kvet and Kropfelova (2005). The history of CWs in the Czech Republic has been reported by Vymazal (2001, 2002). Oxidation–reduction reactions govern microbial and chemical processes occurring in saturated wetland soils and sediments. Oxygen is the major electron acceptor used in redox reactions in aerobic soils. In anaerobic soils (most of the wetlands soils), where O2 is not present, the major electron acceptors are NO3 − , MnO2 , Fe(OH)3 , SO4 2− , and CO2 (Ponnamperuma, 1972; Richardson and Vepraskas, 2000). Theoretically, the electron acceptors are reduced in anaerobic soils in the order shown above, but in reality redox values will vary across soil horizons because organic matter is not uniformly distributed in soils (Richardson and Vepraskas, 2000). Similarly as in wetland soils, redox potential will reflect the relative surplus of electron donors (i.e. organic substances) for oxidation–reduction reactions in wastewaters treated in constructed wetlands. Non-uniform concentrations of pollutants in wastewater will also lead to variable redox readings. Variable pollutant concentrations as well as the flow rate of wastewater are related to the retention time of a CW. Anaerobic conditions are distinguished by a combination of the absence of O2 and a redox potential (EH ) lower than +400 mV (Rowell, 1981). EH values below 400 mV indicate activity of denitrifying bacteria and EH bellow 100 mV reduction of FeIII+ ions, respectively. EH lower than −100 mV indicates reduction of sulphates and organic substances (fermentation) and EH below −200 mV indicates activity of methanogenic bacteria, respectively. However, if microbial processes are intense, the EH can decrease temporarily to values which are lower than those expected for the redox couple functioning in the system. It is known that in such conditions more than one type of microbial process can take place at the same time. Such a situation was observed by Ottow and Fabig (1985) in an artificial system. Shifts in soil aeration status induce changes in the structure of microbial communities and also in the biochemical pathways of active microorganisms. The measurement of redox potential has widely been used in order to characterize oxidation–reduction conditions of wetland soils (Ponnamperuma, 1972; Richardson and Vepraskas, 2000). The assessment of soil redox potential has been found particularly useful for characterizing the onset of reducing conditions in a wetland soil (substrate), caused by a lack of O2 following saturation by water (Fiedler and Sommer, 2004). Redox potential is a good indicator for the state of the reed bed, which ultimately results from the accumulation of a large number of different influencing parameters (Kayser et al., 2002). Redox potential in waterlogged wetland soils is affected by water level, activity of microorganisms, by presence of plants, and by concentrations of electron acceptors (Ponnamperuma, 1972). Most of studies were focused on long-term fluctuation of redox potential. However, redox conditions in the rhizosphere may change quickly within a very short period due to the ability of wetland plants to ventilate soil by their aerenchyma. Flessa (1994) found a high increase of redox potential in the
rhizosphere of submerged vascular macrophytes induced by light intensity under laboratory conditions. Also Wießner et al. (2005) found daily variations in the redox state of Juncus effusus rhizosphere under laboratory conditions ranging from −200 to +200 mV, driven by daylight. However, there is not information about this phenomenon in situ either in natural or constructed wetlands. Our study is focused on diel and seasonal fluctuation of EH in situ in a treatment bed of a constructed wetland in order to describe redox conditions in vegetated submerged soil. The aims of the current study were: (i) to monitor diel, seasonal and spatial fluctuation of redox potential in a treatment bed of a constructed wetland, and (ii) to identify the factors responsible for the redox potential fluctuation and in consequence, for activities of microbial community affecting treatment efficiencies of the whole system.
2.
Methods
2.1.
Site description
Redox potential was measured in a treatment bed of the constructed wetland (CW) with subsurface horizontal flow located ˇ in Slavoˇsovice, Czech Republic. CW is 20 km east of Ceské ˇ Budejovice at an altitude of 480 m above sea level. The annual average air temperature is 7.9 ◦ C and the average annual sum of precipitation is 634 mm. CW started operating in August 2001 and it is used for treating wastewater (combined sewage and storm water runoff) from the Slavoˇsovice village. The CW is designed for 150 person equivalents (PE), and about 100 PE are connected to the CW at present. Detailed technical parameters of the studied system are summarized in Fig. 1. After pretreatment, the water flows through a splitter chamber, which divides the water between two parallel reed beds. The reed beds are sealed with a natural clay layer which was present on the site. The reed beds are planted with common reed in a gravel substrate (10–20 mm). The wastewater is distributed to the treatment beds through perforated inlet pipes placed horizontally in the inflow zone at a depth of 0.5 m. The pipes are 0.5 m from the front edge of the treatment bed. The same types of pipes are located 0.5 m from the back edge of the treatment bed in the outflow zone. The inflow and outflow zones are filled with coarse stones (50–100 mm). The average wastewater inflow rate was 0.12 ± 0.10 L s−1 (with maximum of 1.0 L s−1 during extremely strong precipitation). Hydraulic retention time ranges from 8 to 16 days (Holcová, 2007).
2.2.
Redox measurements
Redox potential (EH ) was measured using platinum electrodes. The electrodes were inserted in perforated plastic tubes (100 mm in diameter) and installed vertically in two depths (0.2 and 0.5 m) of the treatment bed along a transect laid from inflow to outflow. The transect was situated in the middle of the treatment bed. EH was measured at the distances of 1, 3, 5, 7 and 13 m from the inflow of the treatment bed. Ten platinum electrodes (made by EDT, Turnov, Czech Republic) were individually connected to the high terminals and the reference electrode (Ag/AgCl) was connected to the
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found. The platinum electrode functioning was controlled in the standard ORP solution (K3 [Fe(CN6 )] + K4 [Fe(CN)6 ] in phosphate buffer). The rates of water inflow and outflow were measured with an ultrasonic probe US1000 (Fiedler, Electronics for Ecology, Czech Republic). Temperature was measured using Pt100 sensors. For a detailed evaluation of EH the measuring period (June 2002 to November 2004) was divided into the cold and warm period of particular years. The cold period is defined as the time between November to April and the warm period is defined as the time from May to October.
2.3.
Fig. 1 – Diagram of the study site—constructed wetland ˇ with horizontal subsurface flow in Slavosovice Village, Czech Republic.
low terminals of the datalogger channels (M4216, Fiedler, Electronics for Ecology, Czech Republic). EH was measured continuously from June 2002 to September 2004 every 15 min. Software Most 2.3 (Fiedler, Electronics for Ecology, Czech Republic) was used for datalogger management. MySQL 5.0 (MySQL AB, Sweden) database was used for final data storage and evaluation. The database was managed by the Internet browser using phpMyAdmin 2.10. EH values were corrected relative to the normal H electrode by adding +200 mV (Bochove et al., 2002). The EH values were not corrected to pH 7, because “mixed redox potential” (more redox couples in pore water) was measured and the pH of pore water was very close to 7.00 (pH 6.87 ± 0.25) during the whole studied period. The electrode platinum tips were brushed in order to remove any surface coatings and cleaned in an HCl–HNO3 solution, but evidence of possibly poisoning of the Pt tips reported by Austin and Huddleston (1999) was not
Statistical methods
Statistical operations were carried out using the STATISTICA 6.0 (StatSoft Inc., USA) statistical package and the R statistical package Version 2.6.0 (R Development Core Team, 2008). The Wilcoxon matched pairs test was used to compare redox potential at two depths (Zar, 1998). This nonparametric test (alternative to the t-test for dependent samples) was used because the data did not have normal distribution (Shapiro–Wilk test statistic was used to test for normality) and the redox potential values in different depths were closely correlated. For the relationship analysis standard Pearson correlation coefficient was used (Zar, 1998).
3.
Results
3.1.
Environmental and water parameters
The average air temperature inside reed stand was 14.5 and 1.3 ◦ C during the warm period (May to October) and during the cold period (November to April), respectively. The average temperature of pore water in the reed bed at the 0.2 m depth was 13.4 and 2.7 ◦ C in the warm and the cold period, respectively. Water temperature differed statistically significantly between the two depths (Table 1). The CW system was very efficient in removing organic pollutions (measured as BOD5 82% on average and COD 74.0% on average) as well as suspended solids (52% on average) (Fig. 2). The removal of total nitrogen was 62.6% on average (total nitrogen in inflow 24.5 mg L−1 and in outflow 10.3 mg L−1 on average) and the removal of total phosphorus was 75.4% on average (total phosphorus in inflow 4.9 mg L−1 and in outflow 1.7 mg L−1 on average) (Picek et al., 2003, 2007; Picek and
Table 1 – Temperatures of air and water flowing in the reed bed during (2002–2004) Temperatures (◦ C)
Periods Warm
Mean of air temperature inside the reed stand (0.2 m above surface) Mean and S.D. of water temperature in depth 0.2 m (centre of the reed bed) Mean and S.D. of water temperature in depth 0.5 m (centre of reed bed) Depths differences in temperatures of water (Z) Results of Wilcoxon matched pairs test (p)
14.5 ± 6.6 13.4 ± 3.8 13.1 ± 3.0 4.3 <0.01
Cold 1.3 ± 7.6 2.7 ± 3.6 3.6 ± 2.5 12.2 <0.01
Whole 7.8 ± 9.7 8.8 ± 6.4 9.0 ± 5.5 6.3 <0.01
Are shown mean and standard deviation (S.D.). Warm period = May to October; cold period = November to April; whole = warm and cold periods together.
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Also, EH was more stable near the inflow than further away from the inflow. Using these data, monthly means of EH were calculated (Fig. 4). Monthly means of EH fluctuated widely from −400 to +500 mV at the distances of 1 and 13 m from inflow during the period 2002–2004. The maximum monthly means of EH were found in April and in October, the minimum values in June and in July. At the end of 2002 and at the beginning of 2003 very low values of the EH were recorded at the 0.2 m depth in the 13 m distance. Finally, the means of all EH data were calculated for each measuring site for the whole period, and for the warm and the cold periods separately (Fig. 5). Surprisingly, we got very consistent values: the values almost did not differ for the layer 0.5 m in the distances 1, 3, 5 and 7 m from inflow, whereas EH increased by 250 mV at the distance 13 m from inflow as compared to 7 m from inflow. A similar trend was found in the upper layer. This trend was opposite to patterns of concentrations of pollutants in the pore water across the reed bed (Fig. 2).
3.3.
Fig. 2 – Concentration of carbon characterized as biological oxygen demand (BOD5 open squares) and chemical oxygen demand (CODCr close circles) in wastewater Bed = inflow to the treatment beds (after pretreatment) and pore water across the reed bed (distances from the inflow, m) of the ˇ Slavosovice constructed wetland during the warm and cold periods of the 2002, 2003 and 2004 years. Shown are means with standard deviation as error bars.
Duˇsek, 2003). The concentrations of organic pollutants measured as BOD5 and COD rapidly decreased within 1–2 m from the inflow (Fig. 2). The patterns of nitrogen and phosphorus concentration (data not shown) were similar to those of BOD5 and COD. Removal efficiency of the studied CW system are comparable with other CW systems built in the Czech Republic (Vymazal, 2001, 2002) and also with other systems built in other countries with similar climatic and environmental conditions (Kadlec and Knight, 1996; Vymazal et al., 1998).
3.2.
Diurnal and seasonal fluctuation of EH
EH fluctuated from −400 to +800 mV. An example of EH measured at a distance 13 m from inflow in the upper layer is shown in Fig. 3. EH fluctuated from −400 to +600 mV in the period from May till July 2003 (Fig. 3a). Surprisingly, the fluctuation was very intensive even within 24 h (in the range of 800 mV, Fig. 3d) and this fluctuation seemed to be regular within some periods (Fig. 3c). EH fluctuated less at the distances 1, 3, 5, 7 m from inflow, the fluctuation was also less in the 0.5 m depth than in the 0.2 m depth (data not shown).
Effect of depth on the redox potential
During the warm, cold and whole (data not shown) measuring period, the redox potential was significantly higher (i.e. less negative) at the 0.2 m depth than at the 0.5 m depth (Fig. 5). The highest differences were found at the distance 13 m from inflow. Results of depth differences analysis by Wilcoxon matched pair test are as follows for the warm period: Z(inflow,1m) = 13.51, Z(3m) = 12.37, Z(5m) = 13.31, Z(7m) = 14.71, Z(outflow,13m) = 12.84, for all distances p < 0.01 and for cold period: Z(inflow,1m) = 0.53, Z(3m) = 1.58, Z(5m) = 10.03, Z(7m) = 9.24, Z(outflow,13m) = 6.67. The difference was not significant at the distances of 1 and 3 m while they were significant at further (5, 7 and 13 m) distances at p < 0.01. EH measured at the 0.2 m depth was positively correlated with values of redox potential in 0.5 m depth except the distance of 13 m from inflow. The closest correlation (0.82) was found for the distance of 1 m from inflow during warm period (Table 2).
3.4.
Relation of redox potential to water temperature
EH was closely correlated with the temperature of interstitial water in the porous substrate of the reed bed (Table 3). Significant negative correlations were found for all distances from
Table 2 – Correlations between redox values measured at depths of 0.2 and 0.5 m Period
Distance from inflow zone to the outflow zone of the treatment bed (m) 1
Warm Cold
0.82 0.66
3
5
7
13
0.35 0.54
0.57 0.46
0.77 0.22
0.56 −0.55
Significant correlations at p < 0.05 are in bold. Warm period = May to October, cold period = November to April. Warm and cold periods since 2002–2004.
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ˇ Fig. 3 – Fluctuation of redox potential at the 0.2 m depth in the 13 m distance in the reed bed of Slavosovice constructed wetland and precipitation in the different time scale. (a) Three months period; (b) 1-month period (June 2003); (c) week (19 June to 25 June 2003); (d) 1 day (22 June 2003). Presented data of the redox potential is corrected relative to the normal H electrode, precipitations as gray bars and average wastewater inflow rate (d, 1 day) during warm period with standard deviation (error lines).
the inflow during the warm period (Table 3). At the depth of 0.5 m, the correlation increased from the inflow (−0.40) to the outflow (−0.74). During the cold period, positive correlations were found for the layer 0.2 m in the distances 1, 3, 5, 7 m from inflow. The correlations for 0.5 m depth were weak and statistically not significant except for the distance 13 m from inflow (Table 3).
period, various effects of the wastewater flow rate on the EH were found. Increasing flow rate caused decreasing EH at the 0.2 m depth, while increasing EH at the 0.5 m depth. During the cold period, EH was usually positively correlated with the water flow rate (except three cases).
4.
Discussion
3.5. Effect of wastewater flow rate on the redox potential
4.1.
Diel changes in EH
EH and wastewater inflow rate were in some cases statistically significantly correlated within a day (Table 4). During the warm
Redox potential fluctuated from −400 to +800 mV. The lowest EH values were extremely low, however, EH about −400 mV were measured by other authors too, e.g. Bochove et al. (2002).
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Fig. 4 – Seasonal course of the redox potential in the reed bed at the distances 1 and 13 m from the inflow at two depths 0.2 and 0.5 m, pore water temperature and precipitation since June 2002 to September 2004 was shown.
In our case, these low EH values were recorded for only very short periods. Such low EH values may be detected in situations, when microbial processes are intense. Then the EH can decrease temporarily to values which are lower than those expected for the redox couple functioning in the system (Ottow and Fabig, 1985). On the other hand, the highest EH values recorded by us were about +800 mV. Such EH indicates presence of oxygen in the system (Rowell, 1981). These extremely high values of EH were measured in clean water at the distance of 13 m from inflow in the upper layer 0.2 m, for short periods. Such situation may be attributed to the plants ability to oxygenate reed bed by their roots. EH increased or decreased by several hundreds millivolts within very short period (hours). EH fluctuated strongly within a day during some periods and this fluctuation was reg-
ular with diurnal cycles—it was shown in example of EH data (Fig. 3). Diurnal cycling of EH was documented also by Flessa (1994) and Wießner et al. (2005), who concluded that EH was affected by transport of oxygen through the plant aerenchyma, which correlated with solar radiation. In such cases EH increased with increasing solar radiation. Plants have the ability to oxygenate the soil by their aerenchymatic tissues. The radial loss of oxygen from the roots to the rhizosphere increases redox potential (Armstrong et al., 2000). If plant ventilation were the main process governing the diel dynamics of redox potential, it would be expected that the redox potential would be the highest around noon. However, in our study, the trend was opposite. EH decreased in the morning, remained low during the day and then increased during the night. This phenomenon could be attributed to the effect
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Table 3 – Correlations between redox values at depths 0.2 and 0.5 m and water temperatures in 2003 year Period
Depth (m)
Distance from inflow of the treatment bed (m) 1
3
5
7
13
Warm
0.2 0.5
−0.44 −0.46 −0.44 −0.63 −0.40 −0.65 −0.65 −0.69
−0.55 −0.74
Cold
0.2 0.5
+0.56 +0.08
−0.04 +0.48
+0.58 +0.41 +0.17 −0.13
+0.47 +0.03
Significant correlations at p < 0.05 are in bold. Warm period = May to October, cold period = November to April.
of plant roots to release exudates to the rhizosphere which serve as a substrate for heterotrophic microorganisms (Paul and Clark, 1989; Bertin et al., 2003). This easily decomposable organic substrate accelerates microbial processes and utilization of oxygen and other electron acceptors, so EH decreases in the result. When root exudation is not intense and oxygen is transported to the roots and released to the soil, EH will increase. But if plant growth is intense, and roots exudation too, then EH will decrease because oxygen is used by microbes rapidly for mineralization of root exudates. It is documented in literature that plants stimulate methanogenesis by root exudation and so supporting microorganisms with organic matter (Altor and Mitsch, 2006; Kankaala et al., 2003). Within some periods of the year 2004, we also found increased methane concentrations inside the reed stems in the distances further than 1 m from inflow in our previous study (Picek et al., 2007) although, in total, methane emission was negligible in the distances further than 1 m from inflow. Therefore, redox conditions were deeply anaerobic around plant roots supporting methanogenesis in a treatment bed in the upper 0.2 m layer too. Such a situation could occur in the plant rhizosphere (near the plant roots) in periods when the plants produced high amounts of root exudates and the oxygen transport to the roots was not sufficient to increase the redox potential. The reason of intense diurnal EH fluctuation could be that the study was performed early after Slavoˇsovice CW was put into operation. In the beginning of Slavoˇsovice CW functioning only little organic matter accumulated in a reed bed and
Table 4 – Diurnal courses correlations coefficients between hourly means of EH measured at the 0.2 and 0.5 m depths and wastewater inflow rate in 2003 and 2004 years Period Depth (m)
Distance from inflow zone to the outflow zone of the treatment bed (m) 1
3
5
7
13
Warm
0.2 0.5
−0.726 0.554
−0.755 0.444
−0.768 −0.371
−0.131 −0.145
−0.585 0.382
Cold
0.2 0.5
0.278 0.602
0.660 0.617
−0.212 0.703
0.630 −0.093
0.659 0.711
Significant correlations at p < 0.05 are given in bold. Warm period = May to October, cold period = November to April.
Fig. 5 – Fluctuation of redox potential at the 0.2 m (open circles) and 0.5 m (close circles) depths across the treatment bed (1 m = inflow zone, 13 m = outflow zone) of the ˇ Slavosovice constructed wetland during the warm and cold periods. Statistical significant differences at p < 0.001 probability levels are by three stars marked (non = not statistical significant difference, p > 0.05). Presented redox potential is corrected relative to the normal H electrode.
therefore the reed bed remained mostly mineral. In older constructed wetlands, where organic matter has already accumulated in gravel or in natural submerged organic soils, redox potential is expected to be more stable and more negative. In such soils, oxygen is used by microbes for mineralization of soil organic matter and therefore the conditions are usually more reducing. Moreover, in mineral soils oxygen can diffuse easier through pore water to deeper layers as compared to organic soils. The EH fluctuation was more intense in 2002–2003 and less intense in 2004 in Slavoˇsovice CW (Fig. 4), which supports the above hypothesis. At the end of 2002 and at the beginning of 2003, very low EH values were recorded at the 0.2 m depth at the 13 m distance. This decrease of EH was probably caused by very low temperatures in winter, when the treatment bed was without dense plant cover (treatment bed planted in late August 2002) and the outflow zone was partly frozen down to 0.1–0.2 m depth. This situation was not recorded at the end of 2003 and the beginning of 2004 (Fig. 4). Although the fluctuation of EH was very intense in the upper layer during the warm period, surprising results were obtained when average values of EH were calculated separately for each measuring point within the warm season. The average EH almost did not differ at the distances 1, 3, 5 and 7 m from inflow in the upper layer. Also, EH was consistently higher in the upper layer as compared to the lower layer. The average EH values calculated for the upper layer indicate, that redox potential is buffered mainly by the ferric/ferrous couple. Concentrations of both iron forms were measured in the pore water of the Slavoˇsovice CW treatment bed and
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both ferric and ferrous ions were present in various ratios (Diáková et al., 2006). Redox conditions in the treatment bed of the Slavoˇsovice CW were not deeply anaerobic, which corresponds to the low methane emission from the reed bed measured by Picek et al. (2007). Methane emission was important only in the wastewater distribution zone, whereas the emissions were negligible in the rest of the treatment bed.
4.2.
Relation of EH to organic pollution
In spite of seasonal differences in EH , some general features were noted. EH varied in general across the reed bed, increasing from inflow to outflow. In addition, higher EH was found in the surface layer (0.2 m depth) as compared with the bottom (0.5 m depth). The same type of EH variation was found by Kadlec and Knight (1996) and García et al. (2003). The low values of EH often measured in the inflow zone were associated with high concentrations of organic pollutants (Kadlec and Knight, 1996; Richardson and Vepraskas, 2000), measured as BOD5 and COD. The EH values also corresponded with the ratio of reduced to oxidised iron in the pore water of the treatment reed bed reported by Diáková et al. (2006). A major part of iron was reduced and only less than 8% was oxidized (FeIII+ form) at the distance of 1 m from inflow, while 30% of iron was found in oxidized form at the distance of 10 m from inflow within some periods (Diáková et al., 2006). As is indicated by the gradients of BOD5 and COD (Fig. 2), organic matter was degraded rapidly in the inflow zone of the treatment bed. This accounted for the strongly anaerobic conditions mainly at the 0.5 m depth at the distances of 1–7 m, as indicated by redox values below −120 mV in average. The upper layer (0.2 m) of the substrate was less deficient in oxygen in comparison with the 0.5 m depth. Remains of oxygen were less than 0.8 mg L−1 during the warm period and less than 1.55 mg L−1 during the cold period at the 0.05–0.10 m depth (Zemanová, 2005).
4.3.
Real and apparent effect of temperature
We found close correlations between EH and pore water temperature (Table 3). The redox potential of any oxidation–reduction reaction is directly influenced by temperature according to the Nernst equation (Chang, 2005). However, this effect is fairly small and therefore not so important for the interpretation of EH in wetland soils and sediments. For example, it is about 5.4 mV in the range of 5–25 ◦ C for the Fe3+ /Fe2+ redox couple. This calculation by the Nernst equation is related to single redox couples, but in the wetland soil solutions contain more redox couples and the resulting EH would be a “mixed potential” (Richardson and Vepraskas, 2000). Temperature has a direct effect on the solubility of oxygen in water and so subsequently also on EH . However, temperature also affects EH indirectly through acceleration of microbial processes and plant growth. The close correlation between EH and pore water temperature indicates that EH is affected by other processes such as microbial activity and plant growth, which are strongly influenced by temperature (Kadlec and Reddy, 2001). We found statistically significant correlation between EH and pore water tempera-
ture, but the patterns differed between the warm and the cold periods (Table 3). Increase of temperature during the warm period accelerated biochemical processes including bacteria activity. Moreover, oxygen solubility in water decreases with increasing temperature. As a result, EH decreases. By contrast, statistically significant positive correlations were found at the 0.2 m depth between EH and the pore water temperature during the cold period (Table 3). This indicates that the biochemical processes were limited by low temperature of the pore water. Further effect of temperature on EH is through evaporation and transpiration. When temperature increases, evaporation and plant transpiration is increased too and water retention time is prolonged. Thus the zone where pollution is most intensely removed from water is shortened—it is closer to the inflow. In this zone microbial processes are running intensely and the value of EH is lowered. The rest of the treatment bed may be more oxidized than during colder periods because concentrations of pollutants are lower there and the plants ventilate treatment bed through their aerenchyma.
4.4.
Effect of wastewater flow rate
The effect of temperature may be masked by differences in the flow regime during the warm and cold period. During the warm period, retention time of water and pollutants is longer because less amount of water incoming to the treatment bed in comparison with cold period and by water loss from the treatment bed owing to evaporation and plant transpiration. The concentrations of organic matter and nutrients in wastewater incoming to the treatment bed can be higher in the warm period than in the cold period. Nevertheless, the concentrations of pollutants in the pore water inside treatment bed measured in the cold period are similar or even higher that in the warm period (Fig. 2). High water flow rate markedly reduced retention time in the Slavoˇsovice CW from 16 days during the warm period to 8 days during the cold period (Holcová, 2007). Also, during cold period, wastewater was usually more diluted by storm water and water from melting snow than during warm period. Moreover, during the cold period treatment bed does not loose water by plant transpiration and evaporation is very low. Reduced retention time by high flow rates probably accounted for generally smaller differences in EH between the two depths during the cold period as compared to the warm period. This small depth differences was possibly associated with the effect of dilution/mixing in the 0.5 m depth, where EH increased. By this EH increase and a weak decrease of EH in the 0.2 m by the ice cover forming on the surface of treatment bed, EH was similar in both depths at the 1and 3-m distances (Fig. 5). The further distances of the treatment bed were less influenced by the water flow rate and the effect of mixing was less than in the nearer distances from inflow. Close diurnal correlations between EH and water flow rate (Table 4) at the distances of 1, 3, 5, and 13 m of the treatment bed during the warm period indicate that the greater flow rate (greater organic and nutrients loads) strongly affected the EH . However, the response of the EH was different depending on the depth.
e c o l o g i c a l e n g i n e e r i n g 3 4 ( 2 0 0 8 ) 223–232
Organic matter contained in the wastewater enhanced the activity of bacteria associated with plant rhizosphere at the 0.2 m depth and, consequently, the EH decreased. On the other hand, positive correlations, apparent especially at the 0.5 m depth (without roots and rhizomes), can be attributed to the mixing effect of the incoming wastewater. Incoming wastewater with organic matter and nutrients easily disturbed temporary stabilized EH at this depth, because EH in this depth was low. Low temporary equilibrium of EH is very sensitive for any chemical and physical changes (Ponnamperuma, 1972).
4.5. Advantages and limitations of the application of EH for evaluation of redox processes in wetland soil filters Although the measurement of EH has widely been used to characterize oxidation–reduction conditions in wetland soils (Ponnamperuma, 1972; Richardson and Vepraskas, 2000), its interpretation is associated with a lot of uncertainty in the exact sense of chemical science. In general, EH measured in a wetland soil is a result of the current state of different electrochemical systems that are contained in that particular soil and may, or may not be in chemical equilibrium. Exact chemical determination of all of these electrochemical systems is not possible solely on the basis of the redox potential. In highly reduced wetland soils, the redox reading usually stabilizes quite quickly (Fiedler and Sommer, 2004) but this stabilization is not a real equilibrium of the whole wetland system. Wetlands are extremely dynamic systems (owing to biochemical transformations) and a stabilized redox potential value is “a snapshot” of the current situation.
5.
Conclusion
During more than a 2-year period (27 months) redox potential (EH ) fluctuated from −400 to +800 mV in a CW treatment bed according to data measured in situ continuously (in 15 min interval). The EH fluctuation was the most intense in the upper layer (0.2 m) in the furthest distance from inflow, where EH changed by several hundreds millivolts within very short time (2 h) during some periods. According to the EH values measured, aeration conditions in the upper 0.2 m thick layer of the CW treatment bed changed rapidly from deeply anaerobic (−200 mV) to anoxic or almost oxic (+500 mV) and vice versa. These short-time EH fluctuations were attributed mainly to the plant effect—root exudation of soluble organic substances and reed ability to actively transport oxygen to the soil through its aerenchymatic tissues. When average EH was calculated using data for the warm period of a year, it differed only slightly in the upper layer of the CW treatment bed in the transect inflow–outflow, and it only increased near the outflow. Among the factors affecting EH , concentrations of pollutants, pore water temperature and water flow rate were determined. Vegetated treatment beds of subsurface horizontal flow constructed wetlands filled with mineral substrate can be extremely dynamic systems, in which oxidation–reduction status changes within short time from anaerobic to aerobic and vice versa. Anaerobic conditions do not prevail necessar-
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ily in the upper layer of the treatment bed in which plant roots are present.
Acknowledgements The authors gratefully acknowledge the financial support of research provided by projects 206/03/P021 and 206/06/0058 from the Grant Agency of the Czech Republic, project MSM6007665801 from Ministry of Education. We also warmly thank to Slavoˇsovice and Libín village authority leaded by the mayor Mr. Ludvík and constructed wetland manager Mr. Adamík for technical assistance and help with continuous measurement in the Slavoˇsovice constructed wetland. Thanks to Dr. Svatava Kˇrivancová from the Czech Hydrometeorologˇ ˇ ical Institute in Ceské Budejovice for providing precipitation data. For the excellent English revision of the manuscript and valuable remarks thank to Dr. Keith Edwards.
references
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Riverine Wetlands for Optimal Control of Wastewater at Catchment Scale. Publicationes Instituti Geographici Universitatis Tartuensis 94. University of Tartu, Institute of Geography, pp. 44–47. ˇ zková, ˇ Picek, T., Cí H., Duˇsek, J., 2007. Greenhouse gas emission from a constructed wetland—plants as important source of carbon. Ecol. Eng. 31, 98–106. Ponnamperuma, F.N., 1972. The chemistry of submerged soils. Adv. Agron. 24 (29), 29–96. R Development Core Team, 2008. R: a language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. ISBN 3-900051-07-0. http://www.R-project.org. Richardson, J.l., Vepraskas, M.J., 2000. Wetland Soils: Genesis, Hydrology, Landscapes and Classification. Lewis Publishers, USA, pp. 85–97. Rowell, D.L., 1981. Oxidation and reduction. In: Greenland, D.J., Hayes, M.H.B. (Eds.), The Chemistry of Soil Processes. Wiley, Toronto, pp. 401–462. Vymazal, J., Brix, H., Cooper, P.F., Green, M.B., Haberl, R. (Eds.), 1998. Constructed Wetlands for Waste Water Treatment in Europe. Backhuys Publishers, Leiden, The Netherlands, pp. 95–121. Vymazal, J., 2001. Constructed wetlands for wastewater treatment in the Czech Republic. Water Sci. Technol. 44, 369–374. Vymazal, J., 2002. The use of sub-surface constructed wetlands for wastewater treatment in the Czech Republic: 10 years experience. Ecol. Eng. 18, 633–646. Vymazal, J., Kropfelova, L., 2005. Growth of Phragmites australis and Phalaris arundinacea in constructed wetlands for wastewater treatment in the Czech Republic. Ecol. Eng. 25, 606–621. Wießner, A., Kappelmeyer, U., Kuschk, P., Kästner, M., 2005. Influence of the redox condition dynamics on the removal efficiency of a laboratory-scale constructed wetland. Water Res. 39, 248–256. Zar, J.H., 1998. Biostatistical Analysis. Prentice Hall, USA, p. 929. Zemanová, K., 2005. Microbial transformations of carbon, nitrogen and phosphorus in substrate of constructed wetland. M.Sc. Thesis. Faculty of Biological Sciences, University of ˇ ˇ South Bohemia, Ceské Budejovice, Czech Republic, 70 pp. (in Czech).
available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/ecoleng
Redox potential dynamics in a horizontal subsurface flow constructed wetland for wastewater treatment: Diel, seasonal and spatial fluctuations a,c ˇ zková ˇ a,b,∗ , Tomásˇ Picek b , Hana Cí ˇ Jiˇrí Dusek a b c
ˇ Czech Republic Wetland Ecology, Institute of Systems Biology and Ecology, Tˇrebon, ˇ Department of Ecosystem Biology, University of South Bohemia, Faculty of Science, Ceské Budˇejovice, Czech Republic ˇ Department of Biology, University of South Bohemia, Faculty of Agriculture, Ceské Budˇejovice, Czech Republic
a r t i c l e
i n f o
a b s t r a c t
Article history:
Redox potential was measured in a treatment bed of a constructed wetland (CW) with sub-
Received 20 May 2008
surface horizontal flow used for municipal wastewater treatment. CW was situated in the
Received in revised form
Czech Republic, designed for 150 person equivalents; it was planted with Phragmites aus-
4 August 2008
tralis and put into operation in 2001. The system was very efficient in removing organic
Accepted 5 August 2008
pollution (BOD5 82% and COD 74.0%), nitrogen (62.6%), phosphorus (75.4%) and suspended solids (52%). Redox potential (EH ) was measured in two depths (0.2 and 0.5 m) of the treatment bed at regular distances from inflow using platinum electrodes, continuously, every
Keywords:
15 min during a more-than-2-year period (27 months) from 2002 to 2004. EH ranged from
Redox potential
−400 to +800 mV. The fluctuations were the most intense in the upper layer (0.2 m) at the
Redox processes
furthest distance from inflow (13 m) and EH changed within very short time periods (hours)
Phragmites australis
by several hundred millivolts. Among the factors affecting EH concentrations of pollutants,
Wastewater treatment
pore water temperature and water flow rate were determined.
Constructed wetlands Continuous measurement
Vegetated treatment beds of subsurface horizontal flow constructed wetlands filled with mineral substrate can be extremely dynamic systems, in which oxidation–reduction status changes within short time from anaerobic to aerobic and vice versa. Anaerobic conditions do not prevail necessarily in the upper layer of treatment bed, in which plant roots are present. © 2008 Elsevier B.V. All rights reserved.
1.
Introduction
Constructed wetlands (CWs) are becoming increasingly popular worldwide for removing organic matter, nutrients, trace elements, pathogens, or other pollutants from wastewater and/or runoff. CWs used for wastewater treatment provide a relatively simple and inexpensive solution for the treatment of wastewaters from small communities and industries, as well as storm sewage, and agricultural runoff (Kadlec and Knight, 1996). These treatment systems are engineered systems that
∗
Corresponding author. Tel.: +420 384706182. E-mail address: [email protected] (J. Duˇsek). URL: http://www.usbe.cas.cz/index.php?node=466 (J. Duˇsek).
0925-8574/$ – see front matter © 2008 Elsevier B.V. All rights reserved. doi:10.1016/j.ecoleng.2008.08.008
have been designed and constructed to utilize natural processes (Mitsch and Gosselink, 1986; Brix, 1989). The natural processes involve wetland vegetation, soils, and their associated microbial assemblages to assist in treating wastewater. Most existing CWs in the Czech Republic are systems with subsurface horizontal wastewater flow. These systems are designed to remove organics (BOD5 and COD) and suspended solids as those parameters are usually limiting in the effluent. The systems treat municipal or domestic sewage often combined with storm water runoff. The treatment beds of these
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systems are usually planted with common reed (Phragmites australis Trin. Ex Steudel), reed canary grass (Phalaris arundinacea L.) or with both plant species. The growth of common reed and reed canary grass in the CWs was described by Duˇsek ˇ (2001), Duˇsek et al. (2001), and Vymazal (1997), Duˇsek and Kvet and Kropfelova (2005). The history of CWs in the Czech Republic has been reported by Vymazal (2001, 2002). Oxidation–reduction reactions govern microbial and chemical processes occurring in saturated wetland soils and sediments. Oxygen is the major electron acceptor used in redox reactions in aerobic soils. In anaerobic soils (most of the wetlands soils), where O2 is not present, the major electron acceptors are NO3 − , MnO2 , Fe(OH)3 , SO4 2− , and CO2 (Ponnamperuma, 1972; Richardson and Vepraskas, 2000). Theoretically, the electron acceptors are reduced in anaerobic soils in the order shown above, but in reality redox values will vary across soil horizons because organic matter is not uniformly distributed in soils (Richardson and Vepraskas, 2000). Similarly as in wetland soils, redox potential will reflect the relative surplus of electron donors (i.e. organic substances) for oxidation–reduction reactions in wastewaters treated in constructed wetlands. Non-uniform concentrations of pollutants in wastewater will also lead to variable redox readings. Variable pollutant concentrations as well as the flow rate of wastewater are related to the retention time of a CW. Anaerobic conditions are distinguished by a combination of the absence of O2 and a redox potential (EH ) lower than +400 mV (Rowell, 1981). EH values below 400 mV indicate activity of denitrifying bacteria and EH bellow 100 mV reduction of FeIII+ ions, respectively. EH lower than −100 mV indicates reduction of sulphates and organic substances (fermentation) and EH below −200 mV indicates activity of methanogenic bacteria, respectively. However, if microbial processes are intense, the EH can decrease temporarily to values which are lower than those expected for the redox couple functioning in the system. It is known that in such conditions more than one type of microbial process can take place at the same time. Such a situation was observed by Ottow and Fabig (1985) in an artificial system. Shifts in soil aeration status induce changes in the structure of microbial communities and also in the biochemical pathways of active microorganisms. The measurement of redox potential has widely been used in order to characterize oxidation–reduction conditions of wetland soils (Ponnamperuma, 1972; Richardson and Vepraskas, 2000). The assessment of soil redox potential has been found particularly useful for characterizing the onset of reducing conditions in a wetland soil (substrate), caused by a lack of O2 following saturation by water (Fiedler and Sommer, 2004). Redox potential is a good indicator for the state of the reed bed, which ultimately results from the accumulation of a large number of different influencing parameters (Kayser et al., 2002). Redox potential in waterlogged wetland soils is affected by water level, activity of microorganisms, by presence of plants, and by concentrations of electron acceptors (Ponnamperuma, 1972). Most of studies were focused on long-term fluctuation of redox potential. However, redox conditions in the rhizosphere may change quickly within a very short period due to the ability of wetland plants to ventilate soil by their aerenchyma. Flessa (1994) found a high increase of redox potential in the
rhizosphere of submerged vascular macrophytes induced by light intensity under laboratory conditions. Also Wießner et al. (2005) found daily variations in the redox state of Juncus effusus rhizosphere under laboratory conditions ranging from −200 to +200 mV, driven by daylight. However, there is not information about this phenomenon in situ either in natural or constructed wetlands. Our study is focused on diel and seasonal fluctuation of EH in situ in a treatment bed of a constructed wetland in order to describe redox conditions in vegetated submerged soil. The aims of the current study were: (i) to monitor diel, seasonal and spatial fluctuation of redox potential in a treatment bed of a constructed wetland, and (ii) to identify the factors responsible for the redox potential fluctuation and in consequence, for activities of microbial community affecting treatment efficiencies of the whole system.
2.
Methods
2.1.
Site description
Redox potential was measured in a treatment bed of the constructed wetland (CW) with subsurface horizontal flow located ˇ in Slavoˇsovice, Czech Republic. CW is 20 km east of Ceské ˇ Budejovice at an altitude of 480 m above sea level. The annual average air temperature is 7.9 ◦ C and the average annual sum of precipitation is 634 mm. CW started operating in August 2001 and it is used for treating wastewater (combined sewage and storm water runoff) from the Slavoˇsovice village. The CW is designed for 150 person equivalents (PE), and about 100 PE are connected to the CW at present. Detailed technical parameters of the studied system are summarized in Fig. 1. After pretreatment, the water flows through a splitter chamber, which divides the water between two parallel reed beds. The reed beds are sealed with a natural clay layer which was present on the site. The reed beds are planted with common reed in a gravel substrate (10–20 mm). The wastewater is distributed to the treatment beds through perforated inlet pipes placed horizontally in the inflow zone at a depth of 0.5 m. The pipes are 0.5 m from the front edge of the treatment bed. The same types of pipes are located 0.5 m from the back edge of the treatment bed in the outflow zone. The inflow and outflow zones are filled with coarse stones (50–100 mm). The average wastewater inflow rate was 0.12 ± 0.10 L s−1 (with maximum of 1.0 L s−1 during extremely strong precipitation). Hydraulic retention time ranges from 8 to 16 days (Holcová, 2007).
2.2.
Redox measurements
Redox potential (EH ) was measured using platinum electrodes. The electrodes were inserted in perforated plastic tubes (100 mm in diameter) and installed vertically in two depths (0.2 and 0.5 m) of the treatment bed along a transect laid from inflow to outflow. The transect was situated in the middle of the treatment bed. EH was measured at the distances of 1, 3, 5, 7 and 13 m from the inflow of the treatment bed. Ten platinum electrodes (made by EDT, Turnov, Czech Republic) were individually connected to the high terminals and the reference electrode (Ag/AgCl) was connected to the
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found. The platinum electrode functioning was controlled in the standard ORP solution (K3 [Fe(CN6 )] + K4 [Fe(CN)6 ] in phosphate buffer). The rates of water inflow and outflow were measured with an ultrasonic probe US1000 (Fiedler, Electronics for Ecology, Czech Republic). Temperature was measured using Pt100 sensors. For a detailed evaluation of EH the measuring period (June 2002 to November 2004) was divided into the cold and warm period of particular years. The cold period is defined as the time between November to April and the warm period is defined as the time from May to October.
2.3.
Fig. 1 – Diagram of the study site—constructed wetland ˇ with horizontal subsurface flow in Slavosovice Village, Czech Republic.
low terminals of the datalogger channels (M4216, Fiedler, Electronics for Ecology, Czech Republic). EH was measured continuously from June 2002 to September 2004 every 15 min. Software Most 2.3 (Fiedler, Electronics for Ecology, Czech Republic) was used for datalogger management. MySQL 5.0 (MySQL AB, Sweden) database was used for final data storage and evaluation. The database was managed by the Internet browser using phpMyAdmin 2.10. EH values were corrected relative to the normal H electrode by adding +200 mV (Bochove et al., 2002). The EH values were not corrected to pH 7, because “mixed redox potential” (more redox couples in pore water) was measured and the pH of pore water was very close to 7.00 (pH 6.87 ± 0.25) during the whole studied period. The electrode platinum tips were brushed in order to remove any surface coatings and cleaned in an HCl–HNO3 solution, but evidence of possibly poisoning of the Pt tips reported by Austin and Huddleston (1999) was not
Statistical methods
Statistical operations were carried out using the STATISTICA 6.0 (StatSoft Inc., USA) statistical package and the R statistical package Version 2.6.0 (R Development Core Team, 2008). The Wilcoxon matched pairs test was used to compare redox potential at two depths (Zar, 1998). This nonparametric test (alternative to the t-test for dependent samples) was used because the data did not have normal distribution (Shapiro–Wilk test statistic was used to test for normality) and the redox potential values in different depths were closely correlated. For the relationship analysis standard Pearson correlation coefficient was used (Zar, 1998).
3.
Results
3.1.
Environmental and water parameters
The average air temperature inside reed stand was 14.5 and 1.3 ◦ C during the warm period (May to October) and during the cold period (November to April), respectively. The average temperature of pore water in the reed bed at the 0.2 m depth was 13.4 and 2.7 ◦ C in the warm and the cold period, respectively. Water temperature differed statistically significantly between the two depths (Table 1). The CW system was very efficient in removing organic pollutions (measured as BOD5 82% on average and COD 74.0% on average) as well as suspended solids (52% on average) (Fig. 2). The removal of total nitrogen was 62.6% on average (total nitrogen in inflow 24.5 mg L−1 and in outflow 10.3 mg L−1 on average) and the removal of total phosphorus was 75.4% on average (total phosphorus in inflow 4.9 mg L−1 and in outflow 1.7 mg L−1 on average) (Picek et al., 2003, 2007; Picek and
Table 1 – Temperatures of air and water flowing in the reed bed during (2002–2004) Temperatures (◦ C)
Periods Warm
Mean of air temperature inside the reed stand (0.2 m above surface) Mean and S.D. of water temperature in depth 0.2 m (centre of the reed bed) Mean and S.D. of water temperature in depth 0.5 m (centre of reed bed) Depths differences in temperatures of water (Z) Results of Wilcoxon matched pairs test (p)
14.5 ± 6.6 13.4 ± 3.8 13.1 ± 3.0 4.3 <0.01
Cold 1.3 ± 7.6 2.7 ± 3.6 3.6 ± 2.5 12.2 <0.01
Whole 7.8 ± 9.7 8.8 ± 6.4 9.0 ± 5.5 6.3 <0.01
Are shown mean and standard deviation (S.D.). Warm period = May to October; cold period = November to April; whole = warm and cold periods together.
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Also, EH was more stable near the inflow than further away from the inflow. Using these data, monthly means of EH were calculated (Fig. 4). Monthly means of EH fluctuated widely from −400 to +500 mV at the distances of 1 and 13 m from inflow during the period 2002–2004. The maximum monthly means of EH were found in April and in October, the minimum values in June and in July. At the end of 2002 and at the beginning of 2003 very low values of the EH were recorded at the 0.2 m depth in the 13 m distance. Finally, the means of all EH data were calculated for each measuring site for the whole period, and for the warm and the cold periods separately (Fig. 5). Surprisingly, we got very consistent values: the values almost did not differ for the layer 0.5 m in the distances 1, 3, 5 and 7 m from inflow, whereas EH increased by 250 mV at the distance 13 m from inflow as compared to 7 m from inflow. A similar trend was found in the upper layer. This trend was opposite to patterns of concentrations of pollutants in the pore water across the reed bed (Fig. 2).
3.3.
Fig. 2 – Concentration of carbon characterized as biological oxygen demand (BOD5 open squares) and chemical oxygen demand (CODCr close circles) in wastewater Bed = inflow to the treatment beds (after pretreatment) and pore water across the reed bed (distances from the inflow, m) of the ˇ Slavosovice constructed wetland during the warm and cold periods of the 2002, 2003 and 2004 years. Shown are means with standard deviation as error bars.
Duˇsek, 2003). The concentrations of organic pollutants measured as BOD5 and COD rapidly decreased within 1–2 m from the inflow (Fig. 2). The patterns of nitrogen and phosphorus concentration (data not shown) were similar to those of BOD5 and COD. Removal efficiency of the studied CW system are comparable with other CW systems built in the Czech Republic (Vymazal, 2001, 2002) and also with other systems built in other countries with similar climatic and environmental conditions (Kadlec and Knight, 1996; Vymazal et al., 1998).
3.2.
Diurnal and seasonal fluctuation of EH
EH fluctuated from −400 to +800 mV. An example of EH measured at a distance 13 m from inflow in the upper layer is shown in Fig. 3. EH fluctuated from −400 to +600 mV in the period from May till July 2003 (Fig. 3a). Surprisingly, the fluctuation was very intensive even within 24 h (in the range of 800 mV, Fig. 3d) and this fluctuation seemed to be regular within some periods (Fig. 3c). EH fluctuated less at the distances 1, 3, 5, 7 m from inflow, the fluctuation was also less in the 0.5 m depth than in the 0.2 m depth (data not shown).
Effect of depth on the redox potential
During the warm, cold and whole (data not shown) measuring period, the redox potential was significantly higher (i.e. less negative) at the 0.2 m depth than at the 0.5 m depth (Fig. 5). The highest differences were found at the distance 13 m from inflow. Results of depth differences analysis by Wilcoxon matched pair test are as follows for the warm period: Z(inflow,1m) = 13.51, Z(3m) = 12.37, Z(5m) = 13.31, Z(7m) = 14.71, Z(outflow,13m) = 12.84, for all distances p < 0.01 and for cold period: Z(inflow,1m) = 0.53, Z(3m) = 1.58, Z(5m) = 10.03, Z(7m) = 9.24, Z(outflow,13m) = 6.67. The difference was not significant at the distances of 1 and 3 m while they were significant at further (5, 7 and 13 m) distances at p < 0.01. EH measured at the 0.2 m depth was positively correlated with values of redox potential in 0.5 m depth except the distance of 13 m from inflow. The closest correlation (0.82) was found for the distance of 1 m from inflow during warm period (Table 2).
3.4.
Relation of redox potential to water temperature
EH was closely correlated with the temperature of interstitial water in the porous substrate of the reed bed (Table 3). Significant negative correlations were found for all distances from
Table 2 – Correlations between redox values measured at depths of 0.2 and 0.5 m Period
Distance from inflow zone to the outflow zone of the treatment bed (m) 1
Warm Cold
0.82 0.66
3
5
7
13
0.35 0.54
0.57 0.46
0.77 0.22
0.56 −0.55
Significant correlations at p < 0.05 are in bold. Warm period = May to October, cold period = November to April. Warm and cold periods since 2002–2004.
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ˇ Fig. 3 – Fluctuation of redox potential at the 0.2 m depth in the 13 m distance in the reed bed of Slavosovice constructed wetland and precipitation in the different time scale. (a) Three months period; (b) 1-month period (June 2003); (c) week (19 June to 25 June 2003); (d) 1 day (22 June 2003). Presented data of the redox potential is corrected relative to the normal H electrode, precipitations as gray bars and average wastewater inflow rate (d, 1 day) during warm period with standard deviation (error lines).
the inflow during the warm period (Table 3). At the depth of 0.5 m, the correlation increased from the inflow (−0.40) to the outflow (−0.74). During the cold period, positive correlations were found for the layer 0.2 m in the distances 1, 3, 5, 7 m from inflow. The correlations for 0.5 m depth were weak and statistically not significant except for the distance 13 m from inflow (Table 3).
period, various effects of the wastewater flow rate on the EH were found. Increasing flow rate caused decreasing EH at the 0.2 m depth, while increasing EH at the 0.5 m depth. During the cold period, EH was usually positively correlated with the water flow rate (except three cases).
4.
Discussion
3.5. Effect of wastewater flow rate on the redox potential
4.1.
Diel changes in EH
EH and wastewater inflow rate were in some cases statistically significantly correlated within a day (Table 4). During the warm
Redox potential fluctuated from −400 to +800 mV. The lowest EH values were extremely low, however, EH about −400 mV were measured by other authors too, e.g. Bochove et al. (2002).
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Fig. 4 – Seasonal course of the redox potential in the reed bed at the distances 1 and 13 m from the inflow at two depths 0.2 and 0.5 m, pore water temperature and precipitation since June 2002 to September 2004 was shown.
In our case, these low EH values were recorded for only very short periods. Such low EH values may be detected in situations, when microbial processes are intense. Then the EH can decrease temporarily to values which are lower than those expected for the redox couple functioning in the system (Ottow and Fabig, 1985). On the other hand, the highest EH values recorded by us were about +800 mV. Such EH indicates presence of oxygen in the system (Rowell, 1981). These extremely high values of EH were measured in clean water at the distance of 13 m from inflow in the upper layer 0.2 m, for short periods. Such situation may be attributed to the plants ability to oxygenate reed bed by their roots. EH increased or decreased by several hundreds millivolts within very short period (hours). EH fluctuated strongly within a day during some periods and this fluctuation was reg-
ular with diurnal cycles—it was shown in example of EH data (Fig. 3). Diurnal cycling of EH was documented also by Flessa (1994) and Wießner et al. (2005), who concluded that EH was affected by transport of oxygen through the plant aerenchyma, which correlated with solar radiation. In such cases EH increased with increasing solar radiation. Plants have the ability to oxygenate the soil by their aerenchymatic tissues. The radial loss of oxygen from the roots to the rhizosphere increases redox potential (Armstrong et al., 2000). If plant ventilation were the main process governing the diel dynamics of redox potential, it would be expected that the redox potential would be the highest around noon. However, in our study, the trend was opposite. EH decreased in the morning, remained low during the day and then increased during the night. This phenomenon could be attributed to the effect
e c o l o g i c a l e n g i n e e r i n g 3 4 ( 2 0 0 8 ) 223–232
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Table 3 – Correlations between redox values at depths 0.2 and 0.5 m and water temperatures in 2003 year Period
Depth (m)
Distance from inflow of the treatment bed (m) 1
3
5
7
13
Warm
0.2 0.5
−0.44 −0.46 −0.44 −0.63 −0.40 −0.65 −0.65 −0.69
−0.55 −0.74
Cold
0.2 0.5
+0.56 +0.08
−0.04 +0.48
+0.58 +0.41 +0.17 −0.13
+0.47 +0.03
Significant correlations at p < 0.05 are in bold. Warm period = May to October, cold period = November to April.
of plant roots to release exudates to the rhizosphere which serve as a substrate for heterotrophic microorganisms (Paul and Clark, 1989; Bertin et al., 2003). This easily decomposable organic substrate accelerates microbial processes and utilization of oxygen and other electron acceptors, so EH decreases in the result. When root exudation is not intense and oxygen is transported to the roots and released to the soil, EH will increase. But if plant growth is intense, and roots exudation too, then EH will decrease because oxygen is used by microbes rapidly for mineralization of root exudates. It is documented in literature that plants stimulate methanogenesis by root exudation and so supporting microorganisms with organic matter (Altor and Mitsch, 2006; Kankaala et al., 2003). Within some periods of the year 2004, we also found increased methane concentrations inside the reed stems in the distances further than 1 m from inflow in our previous study (Picek et al., 2007) although, in total, methane emission was negligible in the distances further than 1 m from inflow. Therefore, redox conditions were deeply anaerobic around plant roots supporting methanogenesis in a treatment bed in the upper 0.2 m layer too. Such a situation could occur in the plant rhizosphere (near the plant roots) in periods when the plants produced high amounts of root exudates and the oxygen transport to the roots was not sufficient to increase the redox potential. The reason of intense diurnal EH fluctuation could be that the study was performed early after Slavoˇsovice CW was put into operation. In the beginning of Slavoˇsovice CW functioning only little organic matter accumulated in a reed bed and
Table 4 – Diurnal courses correlations coefficients between hourly means of EH measured at the 0.2 and 0.5 m depths and wastewater inflow rate in 2003 and 2004 years Period Depth (m)
Distance from inflow zone to the outflow zone of the treatment bed (m) 1
3
5
7
13
Warm
0.2 0.5
−0.726 0.554
−0.755 0.444
−0.768 −0.371
−0.131 −0.145
−0.585 0.382
Cold
0.2 0.5
0.278 0.602
0.660 0.617
−0.212 0.703
0.630 −0.093
0.659 0.711
Significant correlations at p < 0.05 are given in bold. Warm period = May to October, cold period = November to April.
Fig. 5 – Fluctuation of redox potential at the 0.2 m (open circles) and 0.5 m (close circles) depths across the treatment bed (1 m = inflow zone, 13 m = outflow zone) of the ˇ Slavosovice constructed wetland during the warm and cold periods. Statistical significant differences at p < 0.001 probability levels are by three stars marked (non = not statistical significant difference, p > 0.05). Presented redox potential is corrected relative to the normal H electrode.
therefore the reed bed remained mostly mineral. In older constructed wetlands, where organic matter has already accumulated in gravel or in natural submerged organic soils, redox potential is expected to be more stable and more negative. In such soils, oxygen is used by microbes for mineralization of soil organic matter and therefore the conditions are usually more reducing. Moreover, in mineral soils oxygen can diffuse easier through pore water to deeper layers as compared to organic soils. The EH fluctuation was more intense in 2002–2003 and less intense in 2004 in Slavoˇsovice CW (Fig. 4), which supports the above hypothesis. At the end of 2002 and at the beginning of 2003, very low EH values were recorded at the 0.2 m depth at the 13 m distance. This decrease of EH was probably caused by very low temperatures in winter, when the treatment bed was without dense plant cover (treatment bed planted in late August 2002) and the outflow zone was partly frozen down to 0.1–0.2 m depth. This situation was not recorded at the end of 2003 and the beginning of 2004 (Fig. 4). Although the fluctuation of EH was very intense in the upper layer during the warm period, surprising results were obtained when average values of EH were calculated separately for each measuring point within the warm season. The average EH almost did not differ at the distances 1, 3, 5 and 7 m from inflow in the upper layer. Also, EH was consistently higher in the upper layer as compared to the lower layer. The average EH values calculated for the upper layer indicate, that redox potential is buffered mainly by the ferric/ferrous couple. Concentrations of both iron forms were measured in the pore water of the Slavoˇsovice CW treatment bed and
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both ferric and ferrous ions were present in various ratios (Diáková et al., 2006). Redox conditions in the treatment bed of the Slavoˇsovice CW were not deeply anaerobic, which corresponds to the low methane emission from the reed bed measured by Picek et al. (2007). Methane emission was important only in the wastewater distribution zone, whereas the emissions were negligible in the rest of the treatment bed.
4.2.
Relation of EH to organic pollution
In spite of seasonal differences in EH , some general features were noted. EH varied in general across the reed bed, increasing from inflow to outflow. In addition, higher EH was found in the surface layer (0.2 m depth) as compared with the bottom (0.5 m depth). The same type of EH variation was found by Kadlec and Knight (1996) and García et al. (2003). The low values of EH often measured in the inflow zone were associated with high concentrations of organic pollutants (Kadlec and Knight, 1996; Richardson and Vepraskas, 2000), measured as BOD5 and COD. The EH values also corresponded with the ratio of reduced to oxidised iron in the pore water of the treatment reed bed reported by Diáková et al. (2006). A major part of iron was reduced and only less than 8% was oxidized (FeIII+ form) at the distance of 1 m from inflow, while 30% of iron was found in oxidized form at the distance of 10 m from inflow within some periods (Diáková et al., 2006). As is indicated by the gradients of BOD5 and COD (Fig. 2), organic matter was degraded rapidly in the inflow zone of the treatment bed. This accounted for the strongly anaerobic conditions mainly at the 0.5 m depth at the distances of 1–7 m, as indicated by redox values below −120 mV in average. The upper layer (0.2 m) of the substrate was less deficient in oxygen in comparison with the 0.5 m depth. Remains of oxygen were less than 0.8 mg L−1 during the warm period and less than 1.55 mg L−1 during the cold period at the 0.05–0.10 m depth (Zemanová, 2005).
4.3.
Real and apparent effect of temperature
We found close correlations between EH and pore water temperature (Table 3). The redox potential of any oxidation–reduction reaction is directly influenced by temperature according to the Nernst equation (Chang, 2005). However, this effect is fairly small and therefore not so important for the interpretation of EH in wetland soils and sediments. For example, it is about 5.4 mV in the range of 5–25 ◦ C for the Fe3+ /Fe2+ redox couple. This calculation by the Nernst equation is related to single redox couples, but in the wetland soil solutions contain more redox couples and the resulting EH would be a “mixed potential” (Richardson and Vepraskas, 2000). Temperature has a direct effect on the solubility of oxygen in water and so subsequently also on EH . However, temperature also affects EH indirectly through acceleration of microbial processes and plant growth. The close correlation between EH and pore water temperature indicates that EH is affected by other processes such as microbial activity and plant growth, which are strongly influenced by temperature (Kadlec and Reddy, 2001). We found statistically significant correlation between EH and pore water tempera-
ture, but the patterns differed between the warm and the cold periods (Table 3). Increase of temperature during the warm period accelerated biochemical processes including bacteria activity. Moreover, oxygen solubility in water decreases with increasing temperature. As a result, EH decreases. By contrast, statistically significant positive correlations were found at the 0.2 m depth between EH and the pore water temperature during the cold period (Table 3). This indicates that the biochemical processes were limited by low temperature of the pore water. Further effect of temperature on EH is through evaporation and transpiration. When temperature increases, evaporation and plant transpiration is increased too and water retention time is prolonged. Thus the zone where pollution is most intensely removed from water is shortened—it is closer to the inflow. In this zone microbial processes are running intensely and the value of EH is lowered. The rest of the treatment bed may be more oxidized than during colder periods because concentrations of pollutants are lower there and the plants ventilate treatment bed through their aerenchyma.
4.4.
Effect of wastewater flow rate
The effect of temperature may be masked by differences in the flow regime during the warm and cold period. During the warm period, retention time of water and pollutants is longer because less amount of water incoming to the treatment bed in comparison with cold period and by water loss from the treatment bed owing to evaporation and plant transpiration. The concentrations of organic matter and nutrients in wastewater incoming to the treatment bed can be higher in the warm period than in the cold period. Nevertheless, the concentrations of pollutants in the pore water inside treatment bed measured in the cold period are similar or even higher that in the warm period (Fig. 2). High water flow rate markedly reduced retention time in the Slavoˇsovice CW from 16 days during the warm period to 8 days during the cold period (Holcová, 2007). Also, during cold period, wastewater was usually more diluted by storm water and water from melting snow than during warm period. Moreover, during the cold period treatment bed does not loose water by plant transpiration and evaporation is very low. Reduced retention time by high flow rates probably accounted for generally smaller differences in EH between the two depths during the cold period as compared to the warm period. This small depth differences was possibly associated with the effect of dilution/mixing in the 0.5 m depth, where EH increased. By this EH increase and a weak decrease of EH in the 0.2 m by the ice cover forming on the surface of treatment bed, EH was similar in both depths at the 1and 3-m distances (Fig. 5). The further distances of the treatment bed were less influenced by the water flow rate and the effect of mixing was less than in the nearer distances from inflow. Close diurnal correlations between EH and water flow rate (Table 4) at the distances of 1, 3, 5, and 13 m of the treatment bed during the warm period indicate that the greater flow rate (greater organic and nutrients loads) strongly affected the EH . However, the response of the EH was different depending on the depth.
e c o l o g i c a l e n g i n e e r i n g 3 4 ( 2 0 0 8 ) 223–232
Organic matter contained in the wastewater enhanced the activity of bacteria associated with plant rhizosphere at the 0.2 m depth and, consequently, the EH decreased. On the other hand, positive correlations, apparent especially at the 0.5 m depth (without roots and rhizomes), can be attributed to the mixing effect of the incoming wastewater. Incoming wastewater with organic matter and nutrients easily disturbed temporary stabilized EH at this depth, because EH in this depth was low. Low temporary equilibrium of EH is very sensitive for any chemical and physical changes (Ponnamperuma, 1972).
4.5. Advantages and limitations of the application of EH for evaluation of redox processes in wetland soil filters Although the measurement of EH has widely been used to characterize oxidation–reduction conditions in wetland soils (Ponnamperuma, 1972; Richardson and Vepraskas, 2000), its interpretation is associated with a lot of uncertainty in the exact sense of chemical science. In general, EH measured in a wetland soil is a result of the current state of different electrochemical systems that are contained in that particular soil and may, or may not be in chemical equilibrium. Exact chemical determination of all of these electrochemical systems is not possible solely on the basis of the redox potential. In highly reduced wetland soils, the redox reading usually stabilizes quite quickly (Fiedler and Sommer, 2004) but this stabilization is not a real equilibrium of the whole wetland system. Wetlands are extremely dynamic systems (owing to biochemical transformations) and a stabilized redox potential value is “a snapshot” of the current situation.
5.
Conclusion
During more than a 2-year period (27 months) redox potential (EH ) fluctuated from −400 to +800 mV in a CW treatment bed according to data measured in situ continuously (in 15 min interval). The EH fluctuation was the most intense in the upper layer (0.2 m) in the furthest distance from inflow, where EH changed by several hundreds millivolts within very short time (2 h) during some periods. According to the EH values measured, aeration conditions in the upper 0.2 m thick layer of the CW treatment bed changed rapidly from deeply anaerobic (−200 mV) to anoxic or almost oxic (+500 mV) and vice versa. These short-time EH fluctuations were attributed mainly to the plant effect—root exudation of soluble organic substances and reed ability to actively transport oxygen to the soil through its aerenchymatic tissues. When average EH was calculated using data for the warm period of a year, it differed only slightly in the upper layer of the CW treatment bed in the transect inflow–outflow, and it only increased near the outflow. Among the factors affecting EH , concentrations of pollutants, pore water temperature and water flow rate were determined. Vegetated treatment beds of subsurface horizontal flow constructed wetlands filled with mineral substrate can be extremely dynamic systems, in which oxidation–reduction status changes within short time from anaerobic to aerobic and vice versa. Anaerobic conditions do not prevail necessar-
231
ily in the upper layer of the treatment bed in which plant roots are present.
Acknowledgements The authors gratefully acknowledge the financial support of research provided by projects 206/03/P021 and 206/06/0058 from the Grant Agency of the Czech Republic, project MSM6007665801 from Ministry of Education. We also warmly thank to Slavoˇsovice and Libín village authority leaded by the mayor Mr. Ludvík and constructed wetland manager Mr. Adamík for technical assistance and help with continuous measurement in the Slavoˇsovice constructed wetland. Thanks to Dr. Svatava Kˇrivancová from the Czech Hydrometeorologˇ ˇ ical Institute in Ceské Budejovice for providing precipitation data. For the excellent English revision of the manuscript and valuable remarks thank to Dr. Keith Edwards.
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