Ecological Engineering 64 (2014) 350–359
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Ecological Engineering journal homepage: www.elsevier.com/locate/ecoleng
Constructed wetlands for boron removal: A review Onur Can Türker a , Jan Vymazal b,∗ , Cengiz Türe c a
Faculty of Science and Letters, Department of Biology, AksarayUniversity, Aksaray, Turkey Czech University of Life Sciences Prague, Faculty of Environmental Sciences, Department of Applied Ecology, Prague, Czech Republic c Faculty of Science, Department of Biology, Anadolu University, Eskis¸ehir, Turkey b
a r t i c l e
i n f o
Article history: Received 30 September 2013 Received in revised form 30 December 2013 Accepted 1 January 2014 Available online 28 January 2014 Keywords: Boron Boron removal Constructed wetlands Macrophytes Wastewater treatment
a b s t r a c t Boron (B) contamination in the environment still increases because of various natural sources and anthropogenic activities. Therefore, the problem of removing B from water becomes a worldwide concern due to its toxicity and chronic effects on plants, animals and human health. This situation has generated increasing interest in the use of several wastewater treatment technologies in order to remove B from contaminated water. Constructed wetlands (CWs) present friendly alternative methods to treat wastewater around the world, and are used for removing various contaminants including metals and metalloids. This paper reviews current knowledge regarding the removal process of B, discusses application of B removal, and identiﬁes critical knowledge study ﬁelds of future and gaps. Despite the fact that the sediment is a major sink for the removal of B, plants can play a signiﬁcant role under favorable environmental conditions. The most important environmental factors that affect B removal in CWs are climatic conditions (e.g. transpiration rates), pH, temperature, solutions composition and competing species, hydraulic retention time and supporting media. Further research is needed on the major removal mechanism of B in CWs, namely the applicability of surface ﬂow system, hybrid systems and vertical ﬂow systems to remove B from wastewaters, the role of microorganism in order to enhance B removal efﬁciency. © 2014 Elsevier B.V. All rights reserved.
1. Introduction The elevating cost of energy in recent years together with operation cost of wastewater treatment have led to a strong interest to ﬁnd alternative treatment strategies to conventional technologies (Tu et al., 2010). Constructed wetlands (CWs) represent an eco-friendly alternative for various types of wastewater around the world (Vymazal, 2009). Constructed wetlands are found in Europe, North America, South and Central America, Australia, New Zealand and Oceania, Africa and Asia (Vymazal and Kröpfelová, 2008). Probably more than 100,000 CWs worldwide currently treat over billion litres of water per day (Kadlec and Wallace, 2009; Zhi and Ji, 2012). With the recent rapid growth in wastewater treatment technologies, the issue of boron (B) treatment by CWs has come under the scientiﬁc spotlight. In recent years, several laboratory and ﬁeld experiments have been carried out to determine how CW systems can be applied for B removal (Ye et al., 2003; Murray-Gulde et al., 2003; Kuyucak and Zimmer, 2004; Gross et al., 2007; Allende et al., 2012; Türker et al., 2013a,b). However, the applicability of the CWs is not yet clearly assessed in view of recent ﬁndings on B removal.
In the current paper, B chemistry and behavior in aquatic environments, general B related problems in the aquatic environment and the current knowledge regarding the applicability of CWs to removal of B are reviewed. Boron removal mechanisms in sediment and the role of plants in CWs are also discussed in the review. Finally, data needed for full understanding of possible use of CWs for B removal are presented. 2. Chemical properties of boron Boron is a metalloid (atomic weight 10.811, solid state 298 K, melting point 2349 K, boiling point 4200 K) in group 13 of periodic table, and it is not found as a free element in hydrosphere and lithosphere. It always binds with oxygen to form both borate minerals (borax, ulexite and colemanite) and orthoboric acid (Gemici et al., 2008; Wolska and Bryjak, 2013). The average B concentration in earth’s crust is 10 mg kg−1 and varies from 1 to 500 mg kg−1 depending on the composition of substrate type (Hilal et al., 2011; Wolska and Bryjak, 2013). 3. Boron related problems in the environment
∗ Corresponding author. E-mail address: [email protected]
(J. Vymazal). 0925-8574/$ – see front matter © 2014 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.ecoleng.2014.01.007
Soil environment is sensitive to pollutants because many bacterial activities and plants uptake are facilitated by dissolved phases of pollutants in the soil. Boron concentration in the soils varies
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between 2 and 100 mg kg−1 , with an average concentration of 30 mg kg−1 (Nable et al., 1997). The high concentrations of B in the soil may be the source of B toxicity effects observed in plants (Camacho-Cristóbal et al., 2008). In the recent years a signiﬁcant increase in the amount of B in soil has led to decreased plant growth as well as crop yield especially in arid or semi-arid regions (Nable et al., 1997; Miwa et al., 2007). Boron is relatively soluble in water and commonly causes environmental problems, especially for surface waters where most of the discarded B will end up (Schoderboeck et al., 2011). The most noteworthy problems are in parts of Chile, Turkey, China, New Zealand and many parts of USA (Powell et al., 1997; Ye et al., 2003; Craw et al., 2006; Allende et al., 2012; Türker et al., 2013a). The high solubility of B minerals in water and its potential to cause teratogenic effects have raised global concerns about this element for drinking waters reserves where most of them ﬂow through B-enriched areas. Therefore, World Healty Organization (WHO) recommended 0.5 mg l−1 standard for B in drinking water because the poor performance of B removal processes in the treatment technology in areas with high natural B concentrations (Hilal et al., 2011). In 2009, according to the latest data from the UK and USA on dietary intakes, the Drinking-Water Quality Committee recommended the B guideline value of 2.4 mg l−1 (Wolska and Bryjak, 2013). Boron toxicity in plants is an important factor that can limit crop yield and the quality of production in agricultural areas even if it is a micronutrient essential for plant growth (Miwa et al., 2007). The typical symptoms of B toxicity in the crop plants are yellowing of the leaf tips, chlorotic and necrotic patches in the margin/older leaves and spots on fruits (Nable et al., 1997). Boron toxicity has been reported to reduce yield for crop plants in as many as 80 countries in the regions with alkaline soils couple with high pH, as well as soils with strong rainfall leaching regimes (Shorrocks, 1997; Lehto et al., 2010). Boron toxicity in plants is mainly dependent on B concentrations in both irrigation water and soil solution and on B tolerance mechanisms in plants. Therefore, the B concentration should not exceed 4 mg l−1 and 10 mg l−1 for irrigation water and soil solution, respectively (Nable et al., 1997). The daily intake of B by animals may vary widely, depending on the different species of animals which showed the effects of B exposure on reduced growth, suppression of male reproductive system function and cutaneous disorders (Ku et al., 1993). Experimental animal studies indicated that the most sensitive endpoints for repeat dose of B are seen on the reproductive system associated with reversible inhibition of spermiation (Scialli et al., 2010). Boron is also trace elements in human’s diet and humans need on average 1.4 mg d−1 of B in normal diet with average dietary intake (C¸öl and C¸öl, 2003). However, people living in areas rich in B and who use drinking water supply from B-rich artesian wells may have higher B exposures (Scialli et al., 2010). Excess of B can also be delivered in the human diet through food such as fruit, vegetables and nuts grown in B rich soils and irrigated or contaminated with high concentration B. C¸öl and C¸öl (2003) indicated that these people can be exposed to chronic B intake thorough both food and water.
4. Boron in water Boron is found as several species in water depending on their concentrations in water. The main factor that control B speciation is pH (Tu et al., 2010). At higher concentrations at high pH (≤10), polynuclear B species such as [B3 O3 (OH)5 ]2− and [B4 O5 (OH)4 ]2− are found in the water. However, at low concentrations, mononuclear species boric acid [B(OH)3 ], borate ions [B(OH)4 − ] or boric oxide (B2 O3 ) would be dominant (Hilal et al., 2011). These B
species are water soluble and as such are found in surface water system, usually river, lake and ground water (Neal et al., 2010; Schoderboeck et al., 2011). The total concentration of B in water is shown below along with the sum of the two species. [B]t = [B]B(OH)3 + [B]B(OH)4 −
The equilibrium chemistry in water between boric oxide and boric acid species could be described: B2 O2 ↔ HBO2 ↔ B(OH)3
In freshwater ecosystems, boric acid accounts for approximately 95% of the dissolved B, whereas the borate ion is approximately 5% (Dotsika et al., 2011). Boric acid is moderately soluble in water and behaves as a very weak Lewis acid. The behavior of boric acid in water systems depends on other parameters such as temperature, pressure, pH and ionic strength (Hilal et al., 2011). Chemical speciation of boric acid varies with acidity according to the ﬂowing equilibrium equation: B(OH)3 + H2 O ↔ B(OH)4 − + H+ ;
pKa = 9.15 at 25 ◦ C
The average B concentration of B in surface water is 0.1 mg l−1 (Davis et al., 2002) but higher concentrations can be found in some areas (Schoderboeck et al., 2011). Boron concentrations in surface water range from 10 g l−1 to 200 mg l−1 (Okay et al., 1985; Wyness et al., 2003; Böcük et al., 2013). In general, B is relatively easily mobilized during the water-rock interactions (Wolska and Bryjak, 2013). Many studies have shown that the potential B source for surface water and shallow ground water systems are: (i) natural B contamination in surface waters (C¸öl and C¸öl, 2003; Wyness et al., 2003; Gemici et al., 2008) (ii) fertilizers, herbicides and insecticides (Waggott, 1969; Davidson and Bassett, 1993; Wyness et al., 2003; Hasenmueller and Criss, 2013), (iii) mine waters and wastes (Okay et al., 1985; Neal et al., 1998; Wyness et al., 2003; Böcük et al., 2013; Türker et al., 2013a), (iv) B-enriched detergents and cleaning products released into surface waters through treated and untreated wastewater (Waggott, 1969; Neal et al., 1998; Wyness et al., 2003; Stueber and Criss, 2005; Hasenmueller and Criss, 2013). The use of perborate products increases B concentrations in the domestic wastewater and little or no B is removed during conventional wastewater treatment process (Barth, 2000). The results of domestic wastewater discharge can contribute to elevated B concentration in aquatic habitats (Waggott, 1969; Vengosh et al., 1994; Barth, 2000; Chetelat and Gaillardet, 2005; Hasenmueller and Criss, 2013). Some industrial wastewaters could be enriched with B with concentrations >1 mg l−1 . The major industrial applications are detergents, soap and cleaning products, ceramic, chemical and fertilizer manufacture (Neal et al., 2010). Turek et al. (2007) reported B concentrations between 63.5 and 76.5 mg l−1 in industrial landﬁll leachate from Poland. In the ceramic industry, the wastewater contains 15 mg l−1 B and 2000 mg l−1 suspended solids (Chong et al., 2009). The boron concentrations of 46.5 mg l−1 in electric utility wastewater were reported by Ye et al. (2003). Irawan et al. (2011) reported boron concentration of 745 ± 15 mg l−1 in polarizer manufacturing facility in Tainan, Taiwan. Also, concentration of B can be elevated in area of B mining (Türker et al., 2013a). The boron concentrations in mine efﬂuent can reach values of 2000 mg l−1 (Türker et al., 2013a,b; Böcük et al., 2013). On the other hand, Murray-Gulde et al. (2003) and Rahman (2009) reported B concentration in oil ﬁeld produced water as 7 and 28 mg l−1 , respectively. These reports comes from literature have showed that CWs could have considerable potential to remove B from both domestic and industrial wastewater.
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Table 1 Boron removal efﬁciency of various treatment technologies. Methods
Type of water
B removal efﬁciency
Aqueos solution Aqueos solution
Up to 50% Up to 95%
Öztürk and Kavak (2005) Kıpc¸ak and Özdemir (2012)
Adsorption with active carbon Coagulation
Synthetic wastewater Drinking water Synthetic wastewater
Up to 90% Up to 28% 24%
Geothermal water Synthetic wastewater Industrial wastewater Aqueos solution RO permeate
Up to 95% 70% Up to 85% Up to 40% Up to 99%
Up to 90%
Sea water Aqueos solution
43–78% Up to 90%
Maximum removal at pH 2 and 25 ◦ C. The optimum calcinations temperature was 600 ◦ C and the optimum pH was 6.0. High carbon doses (e.g. 25 g l−1 ) needed Typical removal < 10% At optimum conditions (pH 8.0 and aluminum dose of 7.45 g l−1 ) Best removal at pH 8.0 dependence with time The concentrate pH’s were 11.2 and 12 Membranes applied: AMX–CMX (Neosepta) pH of produced waters <4.5 for 600 bed volumes High percent removal of B was obtained when the particle size of the resin is decreased. Survey of 8 operating RO plants Permeate ﬂow (m3 day−1 ) was 23.0–45.5
Electrodialysis Ion exchange
Consequently, it can be emphasized that the B concentration in aquatic environments still increases. Therefore, the efﬁcient wastewater treatment technology for B removal should be available in order to protect aquatic habitats. 5. Boron removal technologies At present, there is no treatment technology which would be speciﬁcally designed for B removal. Therefore B is always removed in conjuction with the other target parameters. The treatment methods used for the removal of B include coagulation–electrocoagulation, adsorbsion on oxides, adsorbsion on active carbon, ion exchance with basic exhangers, electrodialysis, membrane ﬁltration after complexation, use of B selective resins, with diols as B complexing agents (Table 1). The removal of B through the use of adsorption process and ion exchange systems seems to be the most effective treatment methods. Unfortunately, conventional processes such as sedimentation, coagulation or adsorption remove little or no B from wastewater with low B concentration such as sewage (Hasenmueller and Criss, 2013). 6. Boron removal in constructed wetlands Until now, many authors have documented efﬁcient removal of various contaminants including metals or metalloids in CWs (Vymazal and Kröpfelová, 2008; Kadlec and Wallace, 2009; Marchand et al., 2010; Wu et al., 2013; Vymazal, 2013). The information on the removal of B in various types of CWs is limited (Table 2). Also, the removal processes in CWs which are responsible for B removal have not been understood clearly. The chemistry of B differs from that of other trace elements, and the overall B removal process in CWs is very complex, making the identiﬁcation of speciﬁc removal pathways more difﬁcult. Nevertheless, several experiments have been carried out to determine B removal pathways in CW systems (Ye et al., 2003; Kuyucak and Zimmer, 2004; Gross et al., 2007; Allende et al., 2012; Türker et al., 2013a,b). So far, the studies have indicated that the processes responsible for the removal of B in CWs are sorption and plant uptake. However, environmental factors such as pH, transpiration rate, temperature, solution composition and competing species, hydraulic retention time, ﬁltration media and operational factors are also important factors in B removal process in CWs (Ye et al., 2003; Kröpfelová et al., 2009; Türker et al., 2013a,b). Despite the fact that several studies have reported information about the use of CWs for B removal, detail information on removal
Choi and Chen (1979) Simonnot et al. (2000) Yilmaz et al. (2007) Yilmaz et al. (2008) Sayiner et al. (2008) Turek et al. (2008) Banasiak and Schäfer (2009) Nadav (1999) Kabay et al. (2007) Magara et al. (1998) Dominguez-Tagle et al. (2011)
processes can be lack direct evidence for speciﬁc B removal rate in CWs (Ahn et al., 2001; Zhang et al., 2010). The majority of data about B removal rate in CWs were obtained from microcosms (Ye et al., 2003; Murray-Gulde et al., 2003; Allende et al., 2012) and mesocosms (Gross et al., 2007; Türker et al., 2013a,b) but some data are also available from the full scale CWs (Ghermandi et al., 2007; Kröpfelová et al., 2009) (Table 2). In CWs, regardless of system size, B removal rates ranged from 0% to 65% according to the selection of design parameter, initial B concentrations, presence of biologically absorbable B form (boric acid) in wastewater and natural climatic conditions. The microcosm studies carried out under the laboratory condition suggested that substantial amount of B was removed from wastewater during the short periods of time (Ye et al., 2003); whereas some experiments reported that no or little B was removed in the microcosm (Murray-Gulde et al., 2003; Allende et al., 2012). Ye et al. (2003) found the B removal of 32% for microcosm CWs with different plants species in USA. Murray-Gulde et al. (2003) tested the performance of a hybrid reverse osmosis–CW in USA, and found that no B was removed from the wastewater in a CW vegetated with Typha latifolia. A study from Australia (Allende et al., 2012) reported that subsurface ﬂow CW microcosm vegetated with Phragmites australis removed little B from the wastewater, with approximately 12.5% removal rate at the end of the 7 weeks in the 12 weeks experiment period. The differences in B removal rates in microcosms may be derived from container edge effects. The microcosm studies are advantageous because of simple and economic operation including the possibility of comparing larger number of replicates (Marchand et al., 2010). However, the use of microcosms for a laboratory experiment is sometimes a disadvantage because microcosms may not adequately simulate the natural conditions. The common phenomenon occurring in the microcosms is decreased growth of plant roots due to the limited spaces in containers, with proportionally small and tiny roots crowded along the inner surface of microcosm. Therefore, microcosms result for B removal must be evaluated carefully due to limited applicability under in the real, full-scale systems. Mesocosm experiments showed more efﬁcient removal rates than microcosm studies. Gross et al. (2007) reported a 65% B removal in a recirculatd vertical ﬂow CW planted with Lactuca sativa during the summer in Israel. Türker et al. (2013a) reported average removal o 32% in the ﬁeld study with a polyculture CW vegetated with Phragmites australis and Typha latifolia in Turkey. Also, Türker et al. (2013b) found that under natural climatic condition in Turkey, 40% removal efﬁciency in a mono-culture CW planted with
Table 2 Boron removal in various types of constructed wetlands. Wastewater type
Type of CW
HLRg (cm d−1 )
Mean B removal (%)
Electric utility wastewater
Coarse Colma sand and organic-based potting medium
Brackishproduced waters Munipical wastewater Munipical wastewater Greywater
Typha latifolia, Scirpus validus, Phragmites australis, Phalaris arundiacea Polypogan monspeliensis, Equisetum hyemale, Salicornia virginica, Thalia dealbata, Iris pseudacorus var. nana, Sagittaria latifolia, Azolla caroliniana, Lemna minor, Eichornia crassipes, Nasturtium ofﬁcinale Typha latifolia, Scirpus californicus,
Peagravel (2–8 mm)
Canada Italy USA
SSF CWb Free Water Surface CW RVF CWc d
Max inﬂow B concentration (mg l−1 )
Min outﬂow B concentration (mg l−1 )
Ye et al. (2003)
Typha latifolia, Phragmites australis n.d
Stone (1 cm)
Organic soil, tuff, limestone Gravel (4–8)
Kröpfelová et al. (2009)
Crush rock (4–8)
Kröpfelová et al. (2009)
Kröpfelová et al. (2009) Allende et al. (2012)
Domestic wastewater acidic wastewater
Phragmites australis, Phalaris arundiacea Phragmites australis, Phalaris arundiacea Phragmites australis
Poly-culture Mesocosm CW
Mine dranaige wastewater Mine dranaige wastewater
Mono-culture Mesocosm CW Mono-culture Mesocosm CW
Phragmites australis, Typha latifolia Phragmites australis, Typha latifolia,
Gravel, Zeolit, Cocopeat, Crushed limestone Gravel, Sand
Turkey a b c d e f g
Kuyucak and Zimmer (2004) Ghermandi et al. (2007) Gross et al. (2007)
Türker et al. (2013a) Türker et al. (2013b) Türker et al. (2013b)
Hybrid reverse osmosis constructed wetland (CW). Sub-surface ﬂow CW. Recycled vertical ﬂow CW. Horizontal ﬂow CW. Vertical ﬂow microcosm CW. Hydraulic retention time. Hydraulic loading rate, n.d.: indicates no data.
Murray-Gulde et al. (2003)
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Table 3 Concentration of boron (mg kg−1 ) in natural and constructed wetlands sediments. Locality
Type of wetland
Concentration (mg kg−1 )
Turkey USA USA Turkey Turkey Turkey Italy
Lake Wetland sediment exposed to boron CW microcosm HF CWa HFCWa Natural wetland sediment River sediment receiving various wastewater
277–1788 2–349 45.6 18–66 11–20 0.09–73 0.5
Kazancı et al. (2006) Powell et al. (1997) Ye et al. (2003) Türker et al. (2013b) Türker et al. (2013a) Köse et al. (2012) Bonanno (2011)
Horizontal ﬂow constructed wetlands.
Typha latifolia, as well as 27% removal efﬁciency for a mono-culture CW planted with Phragmites australis. Although mesocosm experiments provide more realistic conditions and provide a link between natural environment conditions and ﬁeld studies than microcosm, these studies have also some limitations in their applicability. Several studies have demonstrated B removal in large scale CWs. Boron was removed from the wastewater with rates >19% in the system with mono-cultures of Phragmites australis and Typha latifolia from Canada (Kuyucak and Zimmer, 2004). Ghermandi et al. (2007) found that a full scale surface ﬂow CW in Italy decreased B concentration from 461 g l−1 to 427 g l−1 . Kröpfelová et al. (2009) reported that B removal rates ranged from 21% to 25% in three full scales horizontal ﬂow sub-surface CWs from the Czech Republic. On the other hand, Sartaj et al. (1999) reported the maximum B removal rate (91%) in the literature comes from vertical peat ﬁlter for construction debris leachate. 6.1. Boron retention in ﬁltration media Although there are no studies on B speciation in CWs, the results of soil studies have indicated that B exits in the soil mainly as a weak boric acid (H3 BO3 ) (Nable et al., 1997). With increasing pH value, the dissociation is enhanced and borax ions [B(OH)4 − ] are formed in soils. Therefore, it believed that B mostly exits in inorganic forms (boric acid and borax ions) in the wetlands environments. Boron can be retained in sediment or media through binding to organic matter, adsorption to iron oxyhydroxide and ﬁne textured sediments (Powell et al., 1997; Ye et al., 2003; Craw et al., 2006; Bonanno, 2011; Allende et al., 2012; Türker et al., 2013a). Boron is absorbed to organic particles by a ligand exchange mechanism occurring as either a single OH− exchange or as diol or cis-diol complexes (Sartaj and Fernandes, 2006). Boron adsorption capacity of sediment was directly related to organic matter content and showed a positive effect when the organic matter content increased. Powell et al. (1997) reported that the amount of B (ranged from 11 to 176 mg kg−1 B) in a riverine wetland ranged due to the alterations in ﬂow or sediment load. Results from this study indicated that the organic sediments contained more B than the inorganic sediment closer to the river bed. The examples of B concentration in natural and CWs sediments are given in Table 3. Kuyucak and Zimmer (2004) found 47% removal of B using organic peat media in natural treatment facility. Bonanno (2011) noted the low B concentrations in the sediment related to easy leaching of B from organic soils. Despite that, it should be noted that the organic matter holds much of the total B in sediments; thus sorption can be the key mechanism for removal of B both in natural wetlands and CWs. Ye et al. (2003) used a mixture of colma sand and organic potting medium as substrate in microcosm for B removal from electric utility wastewater and found 36% B retention in the wetland media. Gross et al. (2007) reported a 65% B removal in a recycled vertical ﬂow CW ﬁlled with organic soil. Allende et al. (2012) also investigated effects of substrate media on B removal using gravel, limestone, cocopeat and zeolite and reported that the
organic cocopeat substrate was the most effective substrate for B removal, concluding that B mainly binds with organic matter in wetland media. The authors also suggested that other substrates (gravel, limestone, and zeolite) did not substantially removed B as compaed to cocopeat in a short-term experiment. This ﬁnding is in agreement with other previously reported studies. For example, Türker et al. (2013a,b) reported only low B retention in gravel and sand mixture in a CW. It has been suggested that the organic matter content is often positively correlated with B removal in CWs (Allende et al., 2012). Sorption to organic matter may, thus, be the key mechanism for removal of B in CWs. Goldberg (1997, 2004) also suggested that B could be adsorbed by sesquioxides and clay minerals, especially illites and vermiculites. Boron adsorption to clay increases at alkaline conditions, and is relatively low under acid conditions (Craw et al., 2006). Nevertheless, there is a lack of information on adsorption pathways of B in ﬁltration materials in CWs. Oxic soil conditions promote formation of iron oxides, hydroxides and oxyhydroxides which result in metal removal throught adsorption and co-precipitation processes (Marchand et al., 2010). Iron oxidation followed by adsorption of B onto iron oxyhydroxides is considered to be important B-removal mechanism in wetland soils at pH > 4.0 (Craw et al., 2006). Several models have been developed to determine maximum B immobilization and retention capacity over time; among them Freundlich and Langmuir models are the most commonly used for B (Marchand et al., 2010). 6.2. Boron uptake by plants in CWs Boron uptake by plants is still a paradox for scientists (Hu and Brown, 1997). The general opinion is that the uptake of B is a passive process depending on the formation of B complexes within the cell wall and plant water ﬂuxes (Tanaka and Fujiwara, 2008). Boron speciation plays a key role in the uptake mechanism and further translocation. Boron is absorbed from solution by roots mainly as a form of boric acid, and then it may be transported into the aboveground plant parts (Türe and Bell, 2004; Rees et al., 2011). The presence of high concentration of boric acid in the water would have a signiﬁcant effect on B uptake by plants if uptake is solely determined by the free boric acid concentrations in the environment (Hu and Brown, 1997). Boron can be translocated from roots to leaves and stems mainly via xylem mobility, but it has also been shown that phloem plays a role in mobility of B into the plants (Brown and Shelp, 1997). This is in agreement with ﬁndings that absorbed B is primarily transported with vascular system and deposited in the leaves or stems (Rees et al., 2011). Different reasons may explain why B is transported to leaves and stems: (1) functional, in the structure of cell wall and membrane; (2) reproductional, in pollen tube growth, ﬂowering and fruiting, and (3) physiological, in sugar transport and respiration (Loomis and Durst, 1992; Blevins and Lukaszewski, 1998; Nuttall, 2000). Boron uptake by plants is affected by several environmental factors including pH, the type of exchangeable ions present in the
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solution, amount of organic matter in soil and non-soil environments. The pH value is the most important parameter affecting B absorption by plants (Hu and Brown, 1997) and plant uptake increases with decreasing pH because boric acid [B(OH)3 ] predominates in the solution rather than borax [B(OH)4 − ]. The higher transpiration rate may increase B absorption by plants. It has been reported that increased transpiration results in higher B uptake because B uptake by plants in water via vascular systems (Hu and Brown, 1997). Due to the high transpiration rate of wetland plants, some species have been used for B removal from wastewater in CWs (Qian et al., 1999; Türker et al., 2013a,b). Several researches have reported that plants accumulate B in their tissues (Table 4) with Typha latifolia, Phragmites australis and Lemna spp. being the most frequently reported species. In these species, accumulation of B increased as concentrations of B in wastewater increased (Frick, 1985; Sasmaz and Obek, 2009; Böcük et al., 2013; Türker et al., 2013a). Studies carried out with machrophytes have shown that roots accumulate more B than do leaves and stems (Ye et al., 2003; Türker et al., 2013b). In the study of Ye et al. (2003), B concentrations in roots were 2 to 8-fold greater than those oin the shoots, but no further details were provided as to B concentrations of aboveground parts. Türker et al. (2013b) reported that the B concentrations in roots of Typha latifolia and Phragmites australis were within that range of 97 and 106 mg kg−1 , and roots concentrations were about twice higher than in stems and 4 times higher than in the leaves. Böcük et al. (2013) suggest that this poor B translocation from below-ground parts to aboveground parts occurs when plants are saturated with B. On the other hand, Türker et al. (2013a) determined a signiﬁcant translocation for Typha latifolia and Phragmites australis in a poly-culture CW when the wetland expose to B concentrations between 227 and 165 mg l−1 . Moreover, Rees et al. (2011) revealed that B was transported to leaves and stems after being taken up by below-ground parts (roots/rhizomes) in the plants. These ﬁndings support the fact that B is transported from roots to stems and leaves into the xylem with the stream of transpiration water. Thus, harvesting could be a viable option for removing of B from CWs when B is efﬁciently stored into the aboveground biomass. In contrast to emergent macrophytes, ﬂoating plants accumulate B from water environment directly into their biomass. Floating aquatic plants such as Lemna gibba and Lemna minor have been recognized as species with a high potential for B accumulation (Frick, 1985; Marín and Oron, 2007; Sasmaz and Obek, 2009; Böcük et al., 2013). Marín and Oron (2007) proposed to use L. gibba to treat B from wastewater because this plant takes up more B using its whole plant surface than just by the roots of emergent plants. Frick (1985) reported that B amount in L. minor biomass was 1168 mg kg−1 after a 7 day treatment period. Böcük et al. (2013) found that L. gibba accumulated 2711 mg B kg−1 dry weight when the plant was grown in 150 mg l−1 B solution. Boyd (1974) reported that B concentration in L. minor amounted to 2711 mg kg−1 in a Michigan wetland affected by urban pollution. Although some aquatic plants such as Potamogeton spp., Iris spp. and Scirpus spp. have been demonstrated as species with a high potential for storing B in their biomass (Table 4), there is no publication about the use of these species for B phytoremediation even under laboratory conditions.
6.3.1. Effect of pH The pH of water may substantially affect the efﬁciency of B removal in CWs. As mentioned earlier, B is mainly present as boric acid at acidic pH. When pH increases, borate ions concentrations increase and at pH above 9–10, high concentration of OH− is present in the solution. The optimum pH for B sorption is generally reported that in the range of 7–10 (Sartaj and Fernandes, 2006), pH 9–10 was found to be optimum for adsorption on peat soils, pH 9.0 for clay minerals and soils (Meyer, 1992). According to Allende et al. (2012), low pH makes more difﬁcult for B to be removed by sorption on organic matter. However, some have indicated that at the near neutral pH, B can be adsorbed on aluminum and iron hydroxides and kaolinite (Sartaj and Fernandes, 2006; Craw et al., 2006). 6.3.2. Effect of temperature The solubility rates of B compounds decrease as temperature decreases, while the dissociation of boric acid (bioavailable form of B in environment) increases as temperature increase (Tu et al., 2010). Therefore, if B is going to be removed as a boric acid in CWs, high temperature is required. To increase boric acid concentrations in water, water levels can reduce to their lowest mark to obtain elevated water temperature in CWs especially for surface ﬂow CWs during the warmer season (Kadlec and Wallace, 2009). Adsorption of B on sediment increased when temperature increased from 22 to 45 ◦ C (Singh, 1971; Su and Suarez, 1997). However, Sartaj and Fernandes (2006) reported that sorption of B on fresh peat decreased with increasing temperature from 2 to 22 ◦ C. Plants metabolism is more active at high temperature. Hu and Brown (1997) reported that the increase in air temperature increased total B uptake by plants even if the relative humidity was held constant. Similarly, the same researchers emphasized that an increase in root temperature increased B absorption into the plants tissues. Therefore it can be suggested that biochemical processes catalyzed at higher temperatures potentially result in better B removal in CWs.
6.3. The effect of environmental factors on B removal
6.3.3. Effect of water characteristics and competing species in water Several reports have shown that when ionic strength of solution was increased, an increase of B adsorption on substrate was observed (Kistler and Helvacı, 1994). Sartaj and Fernandes (2006) explained this phenomenon with electrical double layer theory thickness of the charged layer on substrate surface decreases with increasing ionic strength of the solution. The effect of B speciation in the plant uptake has been reported by Türker et al. (2013a). These authors suggested that plants take up B from wastewater in their tissues under neutral pH if B was found in the water as boric acid rather than borate ions. The sorption of B is sensitive to the presence of competing compounds such as sulphate and phosphate. Boron adsorption is further limited by competition by sulphate ions in the water. The competing effect of sulphate was negligible when the dissolved sulphate concentration of 2 mg l−1 was similar to that of dissolved B (Craw et al., 2006). Melamed (1993) reported that B mobility was increased with the presence of phosphate in the water. However, it was found that B adsorption by clay soil was not affected by concentrations of silicon, nitrate, sulphate, molybdate and phosphate (Sartaj and Fernandes, 2006). The effect of the presence of competing compounds in removal of B in CWs has not been reported.
Several environmental factors such as pH, temperature, B speciation and seasonal effects can affect the removal of B in CWs. These factors control B removal process and their cumulative effect could be different depending on the operational factors and aqueous solution under the experiment (Lizama et al., 2011).
6.3.4. Effect of climatic conditions and operational factors Experimental results in laboratory conditions mostly do not completely match the results under ﬁeld conditions due to total effects and unpredictable interactions of many variables under outdoor conditions (Böcük et al., 2013). Each test site and natural area
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Table 4 Concentrations of B (mg kg−1 ) in wetland plants growing in natural stands and constructed wetlands. Plant species Constructed wetlands Lemna gibba Lemna gibba Lemna minor Lemna minor Lemna gibba Sagittaria latifolia Phragmites australis Iris pseudacorus Typha latifolia Typha latifolia Phragmites australis Scirpus validus Thalia dealbata Typha latifolia Polypogon monspeliensis Phragmites australis Schoenoplectus tabernaemontani P. australis Natural wetlands Lemna minor Potamogeton pectinatus Potamogeton pectinatus Lemna trisulca Ceratophylum demersum Peltandra virginica Typha latifolia Typha latifolia Typha latifolia Pontenderia cordata Phragmites australis Hydrocharis morsus ranae Acorus calamus Phalaris australis Juncus effusus Potamogeton berchtoldii Potamogeton perfoliatus Stratiotes aloides Sparganium erectum Mentha aquatica Ceratophylum demersum Potamogeton foliosus Typha latifolia Myriophylumspicatum T. angustifolia Carex stricta Glyceria aquatica Scirpus lacustris Nyphaea odorata Nelumbo lutea Myriophylum heterophyllum T. latifolia Eleocharis quadrangulata P. australis Nuphar advena Juncus effusus Hydrocotyle umbellate Utricularia inﬂata Potamogeton diversifolius Typha latifolia Typha domingensis Ceratophylum demersum Scirpus validus Scirpus americanus Panicum hemitomon Glyceria striata Eleocharis equisetoides a b
Data come from
Concentration (mg kg−1 )
Microcosm Microcosm Microcosm Microcosm Mesocosm Microcosm HFCWa Microcosm HFCWa Microcosm Microcosm Microcosm Microcosm HFPCWb Microcosm HFPCWb Mesocosm Full- scale HFCWa
639–2711 900–1900 1168 1168 515–765 >400 356.7 >350 317.7 >200 >200 >200 >200 142.4 >100 39.4 18–23 16.97
Turkey Israel USA USA Turkey USA Turkey USA Turkey USA USA USA USA Turkey USA Turkey USA Czech Republic
Böcük et al. (2013) Marín and Oron (2007) Frick (1985) Frick (1985) Sasmaz and Obek, 2009 Ye et al. (2003) Türker et al. (2013b) Ye et al. (2003) Türker et al. (2013b) Ye et al. (2003) Ye et al. (2003) Ye et al. (2003) Ye et al. (2003) Türker et al. (2013a) Ye et al. (2003) Türker et al. (2013a) Ahn et al., 2001 Vymazal (2009)
Polluted wetland Polluted wetland Polluted wetland Polluted wetland Polluted wetland Wetland exposed to B pollution Wetland exposed to B pollution Polluted wetland Polluted wetland Wetland exposed to boron pollution Polluted wetland Polluted wetland Polluted wetland Polluted wetland Polluted wetland Polluted wetland Natural Wetland Natural Wetland Polluted wetland Polluted wetland Natural Wetland Polluted wetland Natural Wetland Natural Wetland Polluted wetland Polluted wetland Polluted wetland Polluted wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Polluted wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland Natural Wetland
781–2567 52.3–958.5 279.9 190.7 41.7–148 14–129 64–106 5.2–100 92.67 12–86 77.32 38.4–64.2 56.9 54.5 4.6–51 49.6 0.47–43.7 1.06–41.2 40.2 38.3 12.4–29.7 29.3 28.8 16.2–25.6 24.5 21.4 15.0 14.6 11.3 10.9 10.6 10.4 3.7–9.0 8.2 8.2 8.1 7.7 7.6 5.3 5.2 4.6 4.3 3.2 2.7 2.3 2.0 1.2
Michigan Hungary Michigan Hungary Michigan USA USA SE USA Turkey USA Turkey Hungary Germany Italy SE USA Michigan Hungary Hungary Germany Germany Hungary Michigan Wisconsin Wisconsin Germany Germany Germany Germany South Carolina South Carolina South Carolina Wisconsin South Carolina Germany South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina South Carolina
Glandon and McNabb (1978) Kovács et al. (1984) Glandon and McNabb (1978) Kovács et al. (1984) Glandon and McNabb (1978) Powell et al. (1997) Powell et al. (1997) Boyd and Walley (1972) Bocuk (2010) Powell et al. (1997) Bocuk (2010) Kovács et al. (1984) Seidel (1966) Bonanno (2011) Boyd and Walley (1972) Glandon and McNabb (1978) Kovács et al. (1984) Kovács et al. (1984) Seidel (1966) Seidel (1966) Kovács et al. (1984) Glandon and McNabb (1978) Smith et al. (1988) Smith et al. (1988) Seidel (1966) Seidel (1966) Seidel (1966) Seidel (1966) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Smith et al. (1988) Boyd and Walley (1972) Seidel (1966) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972) Boyd and Walley (1972)
Horizontal ﬂow constructed wetland. Horizontal ﬂow polyculture constructed wetland.
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has their own speciﬁc climatic conditions and information about the exact effect of natural climatic conditions on the treatment process is scant and contradictory. Nevertheless, the effects of climatic conditions on B removal can largely be explained by changes in transpiration rate of plants (Türker et al., 2013a). The metabolic activities of plants are more active during summer than in winter and high temperature increases transpiration resulting in an increase in B uptake by plants. Reduced relative humidity (as consequence of differences in vapour pressure deﬁcits) also can increases transpiration, resulting in an increase in B uptake by plants. Oertli (1994) reported that increased duration and intensity of illumination substantially increased B uptake. Türker et al. (2013a) and Türker et al. (2013b) reported B removal of about 30% in a CW that operated during plant vegetation periods. These researches also emphasized that importance of applications carried out under natural outdoor conditions. Homogenous distribution of wastewater into the CWs supply more contact sites between adsorbent and wastewater, suggesting that more B removal may be achieved. Therefore, managers should be careful for the design of distribution of wastewater homogenously into the CWs. 6.3.5. Effect of hydraulic retention time Metal and metalloid removal is a relatively slow process in CWs. Debusk et al. (1996) reported that the range of hydraulic retention times from 8 to 30 days ensured good removal of nutrients and metals. Although there is no speciﬁc research for effect of retention times on B removal, some indirect evidences showed that longer retention time led to additional removal of B in CWs (Ye et al., 2003; Kröpfelová et al., 2009; Türker et al., 2013a,b). Long retention time provides more contact time between the media and plant roots with wastewater which may result in higher B removal by CWs. Kröpfelová et al. (2009) found 25% B removal efﬁciency at a hydraulic retention time of 10 days. Türker et al. (2013a) and Türker et al. (2013b) reported that B can be efﬁciently removed up to 27% from mine wastewater at the retention time of 15 days in the CW. Ye et al. (2003) demonstrated that 32% B removal at 12 days hydraulic retention time. On the other hand, low removal efﬁciency of B has been shown at short hydraulic retention time. For example, Kuyucak and Zimmer (2004) reported only 19% B removal efﬁciency for 3 days retention, and Murray-Gulde et al. (2003) and Ghermandi et al. (2007) found that no B removal was found in a CW with 5 days retention time. Therefore, it has been suggested that removal mechanisms for B proceed very slowly and gradual removal starts only after the period of 5–7 days. There is a need to determine optimum hydraulic retention time for B removal in order to obtain maximum B rermoval. 6.3.6. Effect of media Wetland sediments contain organic substrates that effectively retain B in natural wetlands and peatlands (Powell et al., 1997; Kuyucak and Zimmer, 2004). However, adsorption ability of organic substrate appears to be limited at low pH conditions, as pH is a cruel factor favouring B adsorption in wetlands. The use of organic-based media to improve the removal of B has been suggested by some researchers (Ye et al., 2003; Kuyucak and Zimmer, 2004; Gross et al., 2007; Allende et al., 2012). Ye et al. (2003) suggested that the use of crushed limestone to increase the removal of B due to co-precipitation of B with calcium. Moreover, Evans (1987) found that a solution containing more than 100 mg l−1 CaCl2 and up to 10.8 mg l−1 B can become super saturated and B will co-precipitate with calcium. Conversely, the study of Allende et al. (2012) reported that gravel, limestone and zeolite have a limited B removal capability in CWs for acidic mine drainage wastewater. Türker et al. (2013a) and Türker et al. (2013b) demonstrated that only small
amount of B was retained in gravel and sand mixture in CWs for B mine waters. Therefore, it seems that gravel, limestone and zeolite appears to be ineffective in B removal under the acidic conditions. 6.4. Boron removal pathways in CWs In general, the organic-based sediments are a major sink for B in the natural and CWs (Powell et al., 1997; Ye et al., 2003). Boron can also be sorbed onto the organic matter, as well as iron oxyhydroxide and ﬁne textured such as clay minerals. Depending on environmental conditions, B is taken up by plants mainly in the form of boric acid from wastewater. The most important factor that affects B uptake by plants is transpiration rate. It is possible to obtain high B removal efﬁciency using vegetation in CWs in the summer period under the ﬁeld conditions (Türker et al., 2013a,b). However, it should be noted that these authors used gravel and sand as ﬁltration materials and long hydraulic retention times. These studies also showed that in the absence of B adsorption to organic sediment materials B can primarily be removed by plant uptake under the acidic-neutral pH (Oertli and Grgurevic, 1975). The nature of the co-precipitated B is unknown, but calcium and B can also precipitate at acidic conditions. Therefore, more effective removal of B was measured when crushed limestone was used due to more available surfaces for co-precipitation of B with calcium (Ye et al., 2003). 6.5. Future direction and key research needs The literature reveals that CWs have the potential to remove B from the water but the removal efﬁciency depends on environmental and operational conditions. However, little is known about the B removal mechanisms, and therefore, it is not possible to optimize B removal. It is generally agreed that B can be retained in CWs in the sediment and plants, and the sediments are the major sink for B rather than vegetation. However, more evidences are required to conﬁrm the relevance of the sediment sink for B. Previous studies also showed that plant uptake could play important role in B removal under favorable environmental conditions (Qian et al., 1999; Kuyucak and Zimmer, 2004; Böcük et al., 2013; Türker et al., 2013a,b). Unfortunately, conventional treatment techniques remove little, if any, B from wastewater containing low B concentrations. Boron removal by CWs may, therefore, be a suitable option for the treatment of waters contaminated by B. However, the majority of data describing the removal efﬁciency of B in CWs mostly comes from studies carried out in CWs with sub-surface ﬂow. The efﬁciency of surface ﬂow system, horizontal ﬂow system, hybrid systems, as well as vertical ﬂow systems has not been sufﬁciently studied. Therefore, further research should be aimed at these types of systems and their ability to remove B from wastewaters. Several researches have shown that the microorganisms are key players for the removal of metals in CWs (Kosolapov et al., 2004; Hallberg and Johnson, 2005; Sheoran and Sheoran, 2006; Groudev et al., 2008). However, their roles in the removal of metalloids such as B in CWs are still unknown. A major effect of microorganisms on metal/metalloid removal process under aerobic conditations is the microbial oxidation of Fe and Mn (Murray-Gulde et al., 2005; Lizama et al., 2011). Since the Fe oxyhydroxides can adsorb B (Craw et al., 2006), the activity of iron-oxidising bacteria can be important B removal mechanism in wetlands. Therefore, direct studies should be carried out to evaluate the role of microorganism associated with iron oxidation in the removal of B. It should be noted that organic matter-based substrate and clay minerals are able to sorb B under alkaline conditions. Therefore,
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the use of alternative media to improve the removal of B in CWs will be investigated in future studies. Plants with the highest levels of metal content are usually classiﬁed as hyperaccumulators. To date only a few plants such as Typha latifolia, Phragmites australis, Lemna spp. have been reported as potential hyper accumulator plants for remediation of B from wastewaters. Therefore, phytoremediation potentials of other wetland plant species, especially those occurring in natural wetlands receiving B contamination, are needed to be tested. On the other hand, high concentrations of B in wastewater may inhibit plant growth and thus, may limit the application of CWs for B removal. It can be concluded that B phytotoxicity could lead to a problem with the treatment sustainability. Therefore, the activity of antioxidant enzymes such as catalaze (CAT), glutathione reductase (GR), superoxide dismutase (SOD), ascorbate peroxidase (APX), and guaiacol peroxidase (GPX) should be investigated (Böcük et al., 2013). 7. Conclusions Boron has special chemistry and, thus it mostly differs from that of other trace elements. Boron chemistry depends strongly on pH and ionic strength, and this behavior could be important parameter in B removal process in CWs. The literature on B removal in CWs is very limited, but several studies have showed that CWs have considerable potential to remove B from wastewaters. Nevertheless, there are some limitations to removal of B in CWs: time-consuming process, impact of climatic conditions, long-term sustainability of wetland media and the level of B concentrations in wastewater. According to available data from the literature, the most important B removal mechanisms in the CWs are sorption and plant uptake. The organic-based sediments are a major sink for B removal rather than vegetation in CWs. The main environmental factors that affect B removal in CWs include climatic conditions, transpiration rates, pH, temperature, supporting media texture (e.g. organic matter content), solutions composition and competing species, hydraulic retention time. This review is the ﬁrst to focus on, and integrate available data come from literature regarding B removal in CWs. The review has discussed how CW systems could be applied for B removal and which factors affect B removal process. It is necessary to understand further about: the major removal mechanisms of B in CWs, the applicability of surface ﬂow system, hybrid systems and vertical ﬂow systems to treat B, the role of microorganism should also be investigated. References Ahn, C., Mitsch, W.J., Wolfe, W.E., 2001. Effects of recycled FGD liner material on water quality and macrophytes of constructed wetlands: a mesocosm experiment. Water Res. 35 (3), 633–642. Allende, K.L., Fletcher, T.D., Sun, G., 2012. The effect of substrate media on the removal of arsenic, boron and iron from an acidic wastewater in planted column reactors. Chem. Eng. J. 179, 119–130. Banasiak, L.J., Schäfer, I., 2009. Removal of boron, ﬂuoride and nitrate by electrodialysis in the presence of organic matter. J. Membrane Sci. 334, 101–109. Barth, S.R., 2000. Utilization of boron as a critical parameter in water quality evaluation: implications for thermal and mineral water resources in SW Germany and N Switzerland. Environ. Geol. 40 (1–2), 73–89. Blevins, D.G., Lukaszewski, K.M., 1998. Boron in Plant Structure and Function. Annu. Rev. Plant Physiol. Plant Mol. Biol. 49, 481–500. Bonanno, G., 2011. Trace element accumulation and distribution in the organs of Phragmites australis (common reed) and biomonitoring applications. Ecotoxicol. Environ. Saf. 74, 1057–1064. Boyd, C.E., Walley, W.W., 1972. Studies of the biogeochemistry of boron. I. Concentrations in surface waters, rainfall and aquatic plants. Am. Midl. Nat. 88, 1–14. Bocuk, 2010. Investigation of natural plant diversity on the soils with high boron concentrations in terms of soil-plant relations in west Anatolia. Ph.D. Thesis, Anadolu Univ., Graduate School of Sciences, Eskis¸ehir, Turkey.
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