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Mechanistic insights into the role of river sediment in the attenuation of the herbicide isoproturon
Environmental Pollution 170 (2012) 95e101
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Mechanistic insights into the role of river sediment in the attenuation of the herbicide isoproturon Son B. Trinh 1, Kevin M. Hiscock, Brian J. Reid* School of Environmental Sciences, University of East Anglia, Norwich NR4 7TJ, UK
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 December 2011 Received in revised form 14 May 2012 Accepted 21 May 2012
Mechanistic insights into the relative contribution of sorption and biodegradation on the removal of the herbicide isoproturon (IPU) are reported. 14C-radiorespirometry indicated very low levels of catabolic activity in IPU-undosed and IPU-dosed (0.1, 1, 100 mg L1) river water (RW) and groundwater (GW) (mineralisation: <2%). In contrast, levels of catabolic activity in IPU-undosed and IPU-dosed river sediment (RS) were significantly higher (mineralisation: 14.5e36.9%). Levels of IPU catabolic competence showed a positive log-linear relationship (r2 ¼ 0.768) with IPU concentration present. A threshold IPU concentration of between 0.1 mg L1 and 1 mg L1 was required to significantly (p < 0.05) increase levels of catabolic activity. Given the EU Drinking Water Directive limit for a single pesticide in drinking water of <0.1 mg L1 this result suggests that riverbed sediment infiltration is potentially an appropriate ‘natural’ means of improving water quality in terms of pesticide levels at concentrations that are in keeping with regulatory limits. Ó 2012 Elsevier Ltd. All rights reserved.
Keywords: Isoproturon Sorption Biodegradation Catabolic activity Sediment River water Groundwater Riverbank filtration
1. Introduction The abstraction of GW as a drinking water resource is commonplace throughout the world. Some of this drinking water relies upon bank infiltration to maintain its potability. To this end, undesirable impurities such as dissolved organic carbon (DOC) (Ludwig et al., 1997; Hoppe-Jones et al., 2010), pathogens (Weiss et al., 2004), and herbicides (Verstraeten et al., 2002; Vargha et al., 2005) are removed as water ingresses through the riverbank and is, in effect, ‘filtered’. The capacity of the natural system to fulfil this service is dependent upon a number of factors such as river sediment properties (Younger et al., 1993; Margoum et al., 2006), microbial activity (Warren et al., 2003), contaminant load (Helweg et al., 1998; Ray et al., 2002b), hydrogeological properties (Hiscock and Grischek, 2002) and integration of these factors. Abstraction of drinking water from boreholes adjacent to rivers has been practiced throughout Europe, the United States of America (USA) and in many areas of the developing world (Ray et al., 2002a). The herbicide isoproturon (IPU; 3-(4-isopropylphenyl)-1,1dimethyl urea) is of significant concern on account of its * Corresponding author. E-mail address: [email protected] (B.J. Reid). 1 Present address: Mekong River Commission, 184, Fa Ngum road, Unit18, Sithane Neua, Sikhottabong District, Vientiane 0100, Laos. 0269-7491/$ e see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.envpol.2012.05.026
particularly high usage historically. Monitoring by the Environment Agency of England and Wales in 2006 found that IPU was the commonest cause of breaches of the 0.1 mg L1 European Union limit in surface waters. In Great Britain, more than 2.0 106 kg of IPU were annually applied on more than 1.8 106 treated hectares from 1990 until 2007 (CSL, 2008). A number of studies have revealed the capacity of agricultural soils to respond to IPU input and become catabolically enhanced (Bending et al., 2001; El-Sebai et al., 2005; Reid et al., 2005; El Sebaï et al., 2011). There remains, however, the need to understand the role of RS in acting as a zone of attenuation in riverbank filtration (RBF) schemes and in the processing of diffuse contaminants in the hyporheic zone. The main aims of the present work were to assess the potential for natural attenuation of IPU in aquatic systems. Towards these ends biodegradation capacities in three compartments: RW, GW and RS were made. This research provides understanding of the capacity for RBF to accommodate IPU at a range of concentrations and, by extension, the potential for the application of RBF as a viable mitigation approach to remove herbicide residues from surface water. 2. Materials and methods Full details regarding the collection of environmental materials used in this study, their handling and information regarding their physical and chemical
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properties, along with details regarding reagents used, IPU quantification and statistical methods, are provided in supporting information. Provided below are descriptions of the experiments undertaken: i) a 14C-radiorespirometric study to assess the role of IPU concentration and incubation time on levels of IPU catabolic competence in RW, GW and RS; ii) a fixed-bed column recirculation experiment to establish IPU fate in the presence of the binary pairings of sterile/non-sterile RS and sterile/non-sterile RW, and; iii) the subsequent assessment of catabolic competence, with respect to IPU, in material removed from the fixed-bed columns (after 18 d recirculation). 2.1. Assessment of IPU catabolic competence in RW, GW and RS Assessment of catabolic activity was assessed by 14C-radiorespirometry wherein generation of 14CO2 was used as a tracer of 14C-IPU mineralisation. By monitoring the temporal production of 14CO2 in a test system, the catabolic activity of the microbial communities in RW, GW and RS was determined. The 14 C-radiorespirometry procedure was adapted from that described in Allan et al. (2006). In brief, materials under assessment were contained within respirometry flasks. Flasks were spiked with 14C-IPU (1 kBq delivered in 100 mL of methanol; IPU was uniformly 14C-labelled on its phenyl ring) resulting in a final concentration of 14C-IPU of 0.5 mg L1. A glass scintillation vial (7 mL) containing 1 M NaOH (1 mL) suspended from a crocodile clip was used to absorb 14CO2. This vial was periodically removed and replaced with a fresh vial of 1 M NaOH. Ultima Gold scintillation fluid (6 mL) was added to the removed vials. Following 24 h in darkness, 14C-activity in the removed vial was established by liquid scintillation counting (Canberra Packard Tri-Carb 2900TR). The respirometry assay was undertaken for 30 d from the time of 14C-IPU inoculation. To assess catabolic activity in RW and GW, an aliquot (30 mL, n ¼ 3) was transferred to a respirometry flask. Subsequently, standards of 12C-IPU were spiked into the flasks to give final concentrations of 0 (for IPU-undosed treatments), 0.1, 1.0 and 100 mg L1 (for IPU-dosed treatments). The resultant ‘treatments’ were incubated at room temperature with shaking for up to 30 d prior to the addition of 14C-IPU (to assess levels of catabolic activity). Following incubation times of 0 d, 5 d, 10 d and 30 d, 14C-IPU was spiked to a complete set of treatment flasks (n ¼ 3) to initiate the assessment of catabolic activity. These respirometry assays were undertaken for a subsequent 30 d from the point at which 14C-IPU was added. To assess catabolic activity in RS, portions of RS (10 g) were weighed into respirometry flasks containing sterile distilled water (130 mL, n ¼ 3). These flasks were firstly spiked with 12C-IPU standards (prepared in MiliQ water and delivered in 100 mL) to give concentrations of 0 (for IPU-undosed treatments) 0.1, 1 and 100 mg L1 (for IPU-dosed treatments). The resultant ‘treatments’ were incubated at room temperature (15e20 C) with shaking (as before) for 0 d, 5 d, 10 d and 30 d. After these incubation periods, a volume of 100 mL was removed (the flasks were ‘rested’ without shaking over-night before removing this aliquot). These removed aliquots were subsequently used to determine residual 12C-IPU concentrations using a solid phase extraction (SPE) method (see Supporting Information). The flasks and their residual contents (10 g RS and 30 mL MilliQ water) were then used to assess levels of catabolic activity. Towards these ends, the respirometry flasks were spiked with 14C-IPU and a respirometry assay undertaken for 30 d from the point of 14C-IPU inoculation. 2.2. Fixed-bed column recirculation experiment The fixed-bed column recirculation experiment was set-up as shown in Fig. S1. Four treatments (n ¼ 3) were established using the RW and RS samples in either sterile or unsterilised states: Treatment 1 with both RW and RS sterilised (autoclaved: 121 C for 30 min); Treatment 2 with both RW and RS not sterilized; Treatment 3 with only RS sterilised while RW was not sterilised, and; Treatment 4 with only RW sterilised while RS was not sterilised. After packing RW and RS into the system, an IPU stock solution was spiked in to the glass reservoirs to give a final concentration of 100 mg L1. RW samples were removed from the reservoir periodically over the course of 18 d. The samples were filtered (Millex-GP, 0.22 mm) and kept in a cold room (4 C, darkness) before determining IPU concentrations by HPLC (see Supporting information for analytical details). 2.3. Assessment of IPU catabolic competence in fixed-bed column materials Following 18 d of recirculation, RS in the columns representing Treatments 3 and 4 (these treatments contained, respectively, sterile RS with non-sterile RW and non-sterile RS with sterile RW) were assessed with respect to their levels of IPU catabolic competence using the 14C-radiorespirometry method. Samples (10 g) were removed under aseptic conditions from the columns and placed in respirometry flasks containing sterile distilled water (30 mL). Three flasks were prepared from each column (of which n ¼ 3). Flasks were subsequently spiked with 14 C-IPU (see above) and the 14C-radiorespirometry assay undertaken as described above for a total of 10 d.
3. Results and discussion 3.1. IPU catabolic competence in riverbank materials 3.1.1. IPU catabolic competence in RW and GW Biodegradation of IPU in surface water and GW have received little attention previously. The results presented here are significant in that they indicate the IPU-undosed catabolic activity with regards to IPU in both RW and GW to be very low (<2% IPU mineralisation) (Fig. 1). Additionally, no significant (p < 0.05) changes in levels of catabolic activity in RW and GW were observed following IPU addition (at concentrations of 0.1, 1 and 100 mg L1) and following incubation for up to 30 d (Fig. 1). While Johnson et al. (1998) reported very limited decomposition of IPU in GW. To the best of our knowledge catabolic competence with respect to IPU in RW has not been reported to date.
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Incubation time (d) Fig. 1. Mineralisation extents following addition of 14C-IPU and 30 d of subsequent assay time in RW (A), GW (B) and RS (C) in un-dosed controls (black) and IPU-dosed treatments with IPU concentrations 0.1 mg L1 (white), 1 mg L1 (hatched) and 100 mg L1 (cross hatched). Errors are shown as standard errors (n ¼ 3). Note y-axis scales differ.
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3.1.2. IPU catabolic competence in RS Levels of catabolic activity in un-dosed RS were markedly higher (14.5 1.6%) than those in RW and GW (<2% IPU mineralisation) (Figs. 1 and 2). Higher levels are probably a result of the presence of competent microorganisms. Fries et al. (2008) reported the number of microorganisms (per g) in RS to be 100e1000 times greater than that in RW (per mL). In addition to increased cell numbers, the
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opportunity for microbial attachment on to a solid matrix (Warren et al., 2003) may have assisted in the proliferation of microbial populations where RS was present (these populations typically being associated with surfaces in the environment, rather than living free in solution (Warren et al., 2003)). It is noted that IPU degradation in non-sterile GW has been reported to be negligible in the absence of a solid matrix (Johnson et al., 1998). Mineralisation was observed to develop in 3 stages (Fig. 2). The first stage represented a lag period (in this study, a lag period is defined as the time elapsed between 14C-IPU addition and mineralisation exceeding 5%). Following this lag period, a second stage of marked mineralisation was observed. Finally, mineralisation was observed to plateau. Several previous studies on the degradation of IPU in different agricultural soils have reported that the extent of IPU mineralisation in the treatments with previous exposure to IPU to be significantly higher than that in the treatments without previous IPU exposure (Walker and Welch, 1992; Cox et al., 1996; El-Sebai et al., 2005; Reid et al., 2005).
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Assay time (d) Fig. 2. IPU spike after treatments (6). Errors
catabolic activity in river sediment (RS) following addition of a 14C-IPU 0 d (A), 5 d (B), 10 d (C) and 30 d (D) in undosed (C) and IPU-dosed with IPU concentrations of 0.1 mg L1 (B), 1 mg L1 (,) and 100 mg L1 are shown as standard errors (n ¼ 3).
3.1.3. Extent and rate of IPU mineralisation in RS With the exception of the 30 d incubation, all of the 0.1 mg L1 IPU concentration treatments had significantly (p < 0.05) lower extents of mineralisation compared with the un-dosed control at corresponding incubation periods. In contrast, where IPU concentrations were elevated to 1 mg L1, extents of mineralisation were observed to be significantly higher when compared to the control values at all incubation times (Figs. 1 and 2). Thus, catabolic activity was significantly (p < 0.05) enhanced following IPU addition but subsequently diminished with increasing incubation time. This change is attributed to temporal changes in IPU concentration (see below). Where IPU concentrations were elevated still further, to 100 mg L1, extents of mineralisation were again observed to be significantly (p < 0.05) higher when compared to the control values at all incubation times (Figs. 1 and 2). Furthermore, extents of mineralisation in the 100 mg L1 IPU-dosed treatments were also significantly greater than in the 1 mg L1 IPUdosed treatments at all corresponding incubation times (Figs. 1 and 2). Again, as was the case in the 1 and 100 mg L1 IPU-dosed treatments, the catabolic activity initially increased and then decreased in response to temporal changes in substrate level (see below). Maximum 14C-IPU mineralisation (36.9 2.7%) was observed in the 100 mg L1 IPU-dosed treatments after 5 d. It should be borne in mind that some of the IPU associated 14C will be incorporated into cellular biomass and as a consequence complete (100%) mineralisation is unlikely to be achieved (Macleod and Semple, 2006). Maximum rates of mineralisation were observed in the 100 mg L1 IPU-dosed treatments following 5 d incubation (4.74% d1) (Fig. S2). Comparable maximum rates of mineralisation in IPU-exposed agricultural soil (5.04% d1) have also been reported (Reid et al., 2005). Correlation between extent of mineralisation and the maximum rate of mineralisation revealed a strong linear relationship between these parameters (r2 ¼ 0.81; Fig. 3). This relationship suggests that both the extent of mineralisation and rate of mineralisation are equally valid proxies with which to define catabolic competence. To our knowledge, the reporting of these results is the first instance of IPU mineralisation potential being established for RS. Mineralisation of another phenylurea herbicide, diuron, was recently reported in RS microcosms (Pesce et al., 2010). The extent of diuron mineralisation was however found to be less than 5% after 30 d (varying from 5 to 10 % after 84 d) in RS not exposed to diuron, but, relatively high, up to 50% after 84 d, in the diuron-exposed treatments (Pesce et al., 2010). Results presented here and those
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Catabolic activity ( C-IPU mineralized, %)
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Mineralization rate (% 14CO2 d-1) Fig. 3. Relationship between extent of mineralisation and maximum rate of mineralisation in IPU un-dosed controls (C) and IPU dosed treatments at concentrations of 0.1 (B), 1 (,) and 100 mg L1 (6). Errors are shown as standard errors (n ¼ 3).
of Pesce et al. (2010) provide consistent evidence of microbial catabolic competence to degrade phenylurea herbicides present in RS; and, that this catabolic ability can be enhanced in the presence of these herbicides. Comparison of these results for RS with laboratory microcosm experiments with agricultural soils and aquifer materials provides further context. Several authors have reported 14 C-IPU mineralisation, typically of the order of 5e25%, over assay times of 60e90 d at about 20 C in un-dosed soils (Kubiak et al., 1995; Lehr et al., 1996; Pieuchot et al., 1996; Larsen et al., 2000; Scheunert and Reuter, 2000) with 30e50% 14C-IPU being mineralised within 30 d at 15e20 C being reported for pre-exposed soils (Bending et al., 2001; Sorensen and Aamand, 2001; Sorensen et al., 2003). Similarly, Pesce et al. (2009) reported extensive mineralisation of diuron: up to 35.6% (after 109 days) in a vineyard site previously treated with diuron and >40% (after 21 days) in river sediment samples collected downstream of the vineyard site. Collectively, the evidence presented here (for IPU) and that reported by Pesce et al. (2010) (for diuron) indicated catabolic competence for phenylurea herbicides to be comparable in RS and agricultural soil. 3.1.4. Influence of residual 12C-IPU concentrations on IPU catabolic competence in RS The residual concentrations of IPU in dosed RS treatments decreased over the 30 d incubation period at all treatment concentrations. In the 0.1 mg L1 IPU treatment concentrations decreased as follows: 0.04 0.01 mg L1 (0 d), 0.03 0.01 mg L1 (5 d), 0.03 0.01 mg L1 (10 d), and below detection limit (30 d) (the detection limit was 1 mg L1, however, samples were concentrated by 100 times by SPE before injection on to the HPLC). IPU residual concentration also decreased to below detection limit in the 1 mg L1 treatments; changing as follows: 0.92 0.01 (0 d), 0.61 0.07 (5 d), 0.48 0.05 (10 d) mg L1, and below detection limit (30 d). Although detectable IPU concentrations remained in the 100 mg L1 treatments after 30 d of incubation significant (p < 0.05) successive decreases in IPU concentrations were observed over the incubation period; changing as follows: 86.20 0.86 mg L1 (0 d), 76.31 2.10 mg L1 (5 d), 57.79 2.72 mg L1 (10 d) and 31.31 4.42 mg L1 (30 d). Remaining concentration of IPU presented in the treatment incubations at the time of 14C-IPU addition (i.e. the point at which catabolic competence assays were initiated) was observed to have a strong (r2 ¼ 0.768) log-linear relationship with levels of catabolic activity (Fig. 4). This evidence supports the perspective that levels
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Solution phase IPU concentration (µg L ) Fig. 4. Relationship between the extent of IPU mineralisation and the concentrations of IPU present in the IPU-dosed treatments at the time of 14C-IPU spiking in the 0.1 (B), 1 (,), and 100 mg L1 (6) treatments. Errors are shown as standard errors (n ¼ 3).
of catabolic activity are proportional to concentrations of IPU present. Collectively, the results reported (in Figs. 1, 2 and 4) indicate that a threshold concentration of IPU between 0.1 mg L1 and 1 mg L1 was required to promote a significant increase in catabolic competence. Significantly, the EU Drinking Water Directive (CEC, 1998) limits concentrations of a single pesticide in drinking water to < 0.1 mg L1. Thus, the results reported here suggest that should this drinking water limit be exceeded by a factor of about ten, then the RWeRS system should have the capacity to become catabolically competent and reduce IPU concentrations. In this way we suggest that the RWeRS system is potentially an appropriate means of naturally improving water quality in terms of pesticide levels at concentrations that are in keeping with regulatory limits. It is acknowledged that this conclusion is based upon the results reported here and that these results are inextricably linked to the specific sediment sampled and its exposure history to IPU. It is well recognised that sediment texture and its organic mater content exert considerable control over organic compound partitioning (Chung and Alexander, 1998, 1999). This partitioning, in turn, governing free herbicide concentrations and their influence upon substrate-dependent catabolic activity. The sediment used in this research was of high sandy content (79% sand/20% silt (Table S2)) and of low organic matter content (0.8% O.C. (Table S1)). Thus IPU present would have been of relatively high bioavailability in the treatments. Regarding pre-exposure to IPU, Neal et al. (2000) reported concentrations of several pesticides, including the phenylurea herbicides IPU and diuron, in the Thames River at an areas upstream of the sampling site. IPU concentrations were reported to increase at higher flows while diuron concentrations showed no association with flow rate. The average concentration of IPU in river water was 0.16 mg L1 while the maximum IPU concentration was 1.63 mg L1 (Neal et al., 2000). It is noted that this average value is close to that at which no enhancement in catabolic activity was observed in the respirometer assays (Fig. 2) while the maximum concentration is in excess of the concentration required to significantly increase catabolic activity in RS (Fig. 2). 3.2. Sorption and biodegradation of IPU in a RWeRS system Attenuation of IPU was divided into three phases (Fig. 5): Phase I, a ‘sorption phase’ occurred during the first day (consistent for all treatment types); Phase II, a lag or ‘adaptation phase’ (observed in
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-1 IPU concentration in RW (µg L )
90 80 70 60 Treatment 1 - Sterile RW and RS Treatment 2 - Both non-sterile Treatment 3 - Sterile RS (non-sterile RW) Treatment 4 - Non-sterile RS (sterile RW)
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Fig. 5. Attenuation of IPU in a RWeRS system with sterile and non-sterile RW and RS, error bars present standard error of three replicates. RW: river water; RS: river sediment.
Treatments 2 and 4 only; these treatments both containing nonsterile RS) with ‘adaptation’ taking place between days 2 and 6, and; Phase III a ‘degradation phase’, (again only observed in Treatments 2 and 4) that took place between days 7 and 12. 3.2.1. Phase I e sorption phase Concentration of IPU in RW was rapidly decreased following the first day in all treatments (Fig. 5): Treatment 1 (15.8 1.1% sorption), Treatment 2 (16.7 0.5% sorption), Treatment 3 (14.9 2.5% sorption) and Treatment 4 (17.8 2.2% sorption). IPU sorption in all four treatments was not significantly different (p > 0.05). This suggests that sterilization did not significantly affect the sorption capacity of the RS. The average IPU sorption across all four treatments was determined to be 16.3 0.7%. This result is consistent with the values 9e37 % IPU sorption onto ditch sediment (Margoum et al., 2006). In addition, a sorption phase was noted to occur during the first recirculation day of the experiment; again, this observation being consistent with the sorption equilibrium period of ‘less than one day’ reported for several herbicides (including IPU) onto ditch sediment (Margoum et al., 2006). IPU concentrations did not significantly decrease (p > 0.05) over the subsequent 5 d of recirculation in any of the treatments, with IPU concentrations in Treatments 1, 2, 3 and 4 having decreased by only a further: 4.7 0.9%, 4.7 2.0%, 2.6 1.0% and 3.3 1.6%, respectively (Fig. 5). Based upon the results observed in the sorption phase, several sorption isotherm parameters of IPU in the RWeRS system were calculated using the data for all four treatment types collectively (see Supporting information), these parameters were: the maximum sorption capacity of the RS (CS,max) ¼ 225 10 mg kg1; solid-water distribution coefficient (KD) ¼ 2.80 0.14 L kg1; organic carbon-normalized partition coefficient (KOC) ¼ 349 18 L kg1; and the retardation factor (RD) ¼ 7.9 0.36. Using a first-order model, the average first order sorption rate constant ksorp (h1) based upon all four treatments was established to be 8.6 103 0.6 103. To our best knowledge, there are no previous reports of sorption characteristics of IPU to RS using a fixed-bed column method. Consequently, the contextualisation of our results is made with respect to the sorption of IPU to soils and other environmentally relevant geosorbents. A considerable range in IPU KD values (L kg1) has been reported: in top soils (0e50 cm) varying from 2.73 to 4.39; in subsoil (50e70 cm) varying from 0.75 to 0.91, and; in chalk (2.1e9.9 m below the surface) varying from 0.02 to 0.46 (Johnson et al., 1998). It is noted that, in the case of sorption of hydrophobic herbicides, the solid-water distribution coefficient (KD) can be
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normalised to the organic-carbon content; yielding the organiccarbon-normalised distribution coefficient (KOC) (Karickhoff, 1984; Rae et al., 1998). The KOC value of IPU in surface soil samples (from 10 to 90 cm below surface with total organic carbon of 0.8%) was reported to be 150 L kg1 (Worrall et al., 1996). In agricultural soils, KOC values of IPU varied from 112 to 138 L kg1 (Worrall et al., 1996; Cooke et al., 2004). In aquifer sediments (depths varying 12e19 m) KOC values varied from 19 to 278 L kg1 (Rae et al., 1998). In sediment from the Kishon River (Israel) KOC values of IPU were reported to be 60.12 and 125.60 L kg1 (Chefetz et al., 2004). In comparison to these studies the KOC values of IPU reported here for RS are higher (349 18 L kg1). In this study, the sorption of IPU on to RS was noted to be similar to that in topsoil and much higher than in subsoil, chalk, and sediment with low TOC content. 3.2.2. Phase II e adaptation phase The concentration of IPU remained unchanged in Treatment 1 (abiotic control in which both RW and RS were sterilised) following initial sorption in the first day of recirculation; there was no significant decrease (p > 0.05) between the IPU concentrations between days 1 and 18 (Fig. 5). Similarly, no significant (p > 0.05) decreases in IPU concentration was observed in Treatment 3 (consisting of sterile RS and non-sterile RW) between days 1 and 18 (Fig. 5). In contrast, IPU loss from Treatments 2 and 4 (both containing non-sterile RS) differed significantly (p < 0.05) from those in Treatments 1 and 3 (both of which contained sterile RS) (Fig. 5). These treatments revealed an ‘adaptation phase’ wherein no significant (p > 0.05) difference in IPU concentrations were observed between days 1 and 5 of recirculation. Thereafter, a ‘degradation phase’ was observed in which IPU loss was rapid and complete (Fig. 5). 3.2.3. Phase III e biodegradation phase From days 7e12, the IPU concentrations in Treatments 2 and 4 decreased from, respectively, 74.2 2.5 mg L1 and 79.4 1.68 mg L1 to below the HPLC limit of detection. Degradation rates of IPU in Treatments 2 and 4 were estimated using a zeroorder model during Phase III only (see Supporting information) and were determined to be, respectively, 13.62 0.17 and 17.72 1.43 mg L1 d1. Half-lives of IPU in Treatments 2 and 4 were, therefore, calculated to be 2.67 0.04 d and 2.47 0.26 d, respectively. Biodegradation rates of IPU in Treatments 2 and 4 were not significantly different (p > 0.05). This suggests that although results from Treatment 2 indicated IPU could be degraded by microorganisms living in both RW and RS environments, results from Treatment 4 confirmed that RS-borne microorganisms were primarily responsible for the degradation of IPU. This finding is consistent with the results from the previous experiment that indicated very low levels of catabolic activity in RW (<2%) but significantly high levels of catabolic activity in RS (>14.5%). This finding supports the conclusion that RS has the potential to strongly influence degradation while RW does not. Owing to the absence of RS studies with which to draw comparisons, comparison is made with other geosorbents. Reported IPU degradation in systems containing: i) a combination of GW and topsoil, and ii) a combination of GW and chalk (3.1e7.2 m below the surface) showed that in both cases IPU was degraded to below detection level within 7e14 d (Johnson et al., 1998). A variation in IPU degradation to 50% of application ranging over 6.5e30 d within samples taken from the same field has been reported (Walker et al., 2001). An IPU 50% degradation (t50) has been reported to vary from 14 d to longer than 65 d in fields previously dosed and undosed with IPU (Bending et al., 2006). The IPU biodegradation rates in a GW-chalk system has
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been reported to vary with half-lives from 111 to 273 d, while the half-lives of IPU in topsoil has been reported to be shorter, from 15 to 34 d (Ministry of Agriculture Fisheries and Food, 2000). Thus, the half-lives reported here for non-sterile RS (of the order of 2.5 d) indicate a considerable capacity for RS to remove IPU. 3.3. Enhancement of IPU catabolic competence in RS samples taken from the recirculation experiment RS removed from the fixed-bed columns following 18 d of recirculation showed marked differences in their levels of IPU catabolic competence (Fig. S3). RS removed from Treatment 3 (comprised of sterile RS and non-sterile RW) indicted low levels of catabolic competence; with the extent of 14C-IPU mineralisation reaching only 1.6 0.1% (after 10 d; Fig. S3). This low level of mineralisation suggests minimal competence of RW-borne microorganisms to degrade IPU. This result is consistent with the finding for Treatment 3 during the recirculation experiments wherein no further IPU loss was observed following the initial sorption phase (Fig. 5). In addition, this result is also consistent with the low IPU level of IPU-undosed catabolic competence in RW (<2% after 30 d; Fig. 1). In contrast, RS removed from the fixed-bed column of Treatment 4 (comprised of sterile RW and non-sterile RS) exhibited high levels of IPU catabolic competence. The maximum extent of mineralisation in samples removed from Treatment 4 was 40.0 0.8% (after 10 d; sufficient time for mineralisation to plateau) (Fig. S3). A maximum 14C-IPU mineralisation rate of 29.4% d1 was observed in the first day of the assay (Fig. S3). This high level of mineralisation and rapid mineralisation rate suggests significant competence of RW-borne microorganisms to completely degrade IPU. While the extent of IPU mineralisation observed in RS samples removed from the recirculation experiment (40.0 0.8%) was similar to that observed in the respirometer assays (reaching a maximum extent of 36.9 2.7% in the 100 mg L1 treatment after 5 d) the rate of mineralisation in samples obtained from the recirculation regime (29.4% d1) was far greater than that observed in the initial screening of IPU concentration influence upon catabolic activity (4.74% d1). This evidence supports the suggestion that the final biodegradation phase of the recirculation experiments was responsible for the rapid loss of IPU in Phase III. 4. Conclusions Based upon the results obtained from both the recirculation experiment and 14C-radiorespirometry assays, it is concluded that IPU degradation in RSeRW systems is most strongly dependent upon microbial competence in RS and not in RW. Thus, while opportunities for IPU degradation are low for RW and GW, opportunities are much higher in RS. Results indicate that a threshold IPU concentration of between 0.1 mg L1 and 1 mg L1 was required to promote significant increases in catabolic competence in this study. This value is consistent with the EU Drinking Water Directive (CEC, 1998) that limits concentrations of a single pesticide in drinking water to < 0.1 mg L1. Results reported here suggest that should this limit be exceeded, by a factor of about ten, then the RWeRS system should have the capacity to become catabolically competent and reduce IPU concentrations. In this way we suggest that the RWeRS system is potentially an appropriate means of naturally improving water quality in terms of pesticide levels at concentrations that are in keeping with regulatory limits. This finding is important in the context of RBF as it supports the use of RBF as an appropriate means of water purification where pesticides in RW are elevated above regulatory limits.
Acknowledgements The authors would like to acknowledge the Program 300 of the Committee of Hochiminh City, Vietnam for funding this research with a studentship to Son B. Trinh. Special thanks to Dr Michael Jones and his colleagues at Thames Water Utilities who helped us to access field sites and collect samples. Appendix A. Supporting information Supporting information related to this article can be found online at http://dx.doi.org/10.1016/j.envpol.2012.05.026. References Allan, I.J., Semple, K.T., Hare, R., Reid, B.J., 2006. Prediction of mono- and polycyclic aromatic hydrocarbon degradation in spiked soils using cyclodextrin extraction. Environmental Pollution 144, 562e571. Bending, G.D., Shaw, E., Walker, A., 2001. Spatial heterogeneity in the metabolism and dynamics of isoproturon degrading microbial communities in soil. Biology and Fertility of Soils 33, 484e489. Bending, G.D., Lincoln, S.D., Edmondson, R.N., 2006. Spatial variation in the degradation rate of the pesticides isoproturon, azoxystrobin and diflufenican in soil and its relationship with chemical and microbial properties. Environmental Pollution 139, 279e287. CEC, 1998. Council Directive 98/8/EC on the Placing on the Market of Biocidal Products. Council of the European Community, Brussels. Chefetz, B., Bilkis, Y.I., Polubesova, T., 2004. Sorption-desorption behavior of triazine and phenylurea herbicides in Kishon river sediments. Water Research 38, 4383e4394. Chung, N.H., Alexander, M., 1998. Differences in sequestration and bioavailability of organic compounds aged in dissimilar soils. Environmental Science and Technology 32, 855e860. Chung, N.H., Alexander, M., 1999. Relationship between nanoporosity and other properties of soil. Soil Science 164, 726e730. Cooke, C.M., Shaw, G., Collins, C.D., 2004. Determination of solid-liquid partition coefficients (Kd) for the herbicides isoproturon and trifluralin in five UK agricultural soils. Environmental Pollution 132, 541e552. Cox, L., Walker, A., Welch, S.J., 1996. Evidence for the accelerated degradation of isoproturon in soils. Pesticide Science 48, 253e260. CSL, 2008. Pesticide Usage Statistics. Central Science Laboratory. see. http:// pusstats.csl.gov.uk/index.cfm. El Sebaï, T., Devers-Lamrani, M., Lagacherie, B., Rouard, N., Soulas, G., MartinLaurent, F., 2011. Isoproturon mineralization in an agricultural soil. Biology and Fertility of Soils 47, 427e435. El-Sebai, T., Lagacherie, B., Cooper, J.F., Soulas, G., Martin-Laurent, F., 2005. Enhanced isoproturon mineralisation in a clay silt loam agricultural soil. Agronomy for Sustainable Development 25, 271e277. Fries, J.S., Characklis, G.W., Noble, T.R., 2008. Sediment-water exchange of Vibrio sp. and fecal indicator bacteria: implications for persistence and transport in the Neuse River Estuary, North Carolina, USA. Water Research 42, 941e950. Helweg, A., Fomsgaard, I.S., Reffstrup, T.K., Sorensen, H., 1998. Degradation of mecoprop and isoproturon in soil influence of initial concentration. International Journal of Environmental Analytical Chemistry 70, 133e148. Hiscock, K.M., Grischek, T., 2002. Attenuation of groundwater pollution by bank filtration. Journal of Hydrology 266, 139e144. Hoppe-Jones, C., Oldham, G., Drewes, J.E., 2010. Attenuation of total organic carbon and unregulated trace organic chemicals in U.S. riverbank filtration systems. Water Research 44, 4643e4659. Johnson, A.C., Hughes, C.D., Williams, R.J., Chilton, P.J., 1998. Potential for aerobic isoproturon biodegradation and sorption in the unsaturated and saturated zones of a chalk aquifer. Journal of Contaminant Hydrology 30, 281e297. Karickhoff, S.W., 1984. Organic pollutant sorption in aquatic systems. Journal of Hydraulic Engineering-Asce 110, 707e735. Kubiak, R., Ellssel, H., Lambert, M., Eichhorn, K.W., 1995. Degradation of isoproturon in soil in relation to changes of microbial biomass and activity in small-scale laboratory and outdoor studies. International Journal of Environmental Analytical Chemistry 59, 123e132. Larsen, L., Sorensen, S.R., Aamand, J., 2000. Mecoprop, isoproturon, and atrazine in and above a sandy aquifer: vertical distribution of mineralization potential. Environmental Science & Technology 34, 2426e2430. Lehr, S., Glassgen, W.E., Sandermann, H., Beese, F., Scheunert, I., 1996. Metabolism of isoproturon in soils originating from different agricultural management systems and in cultures of isolated soil bacteria. International Journal of Environmental Analytical Chemistry 65, 231e243. Ludwig, U., Grischek, T., Nestler, W., Neumann, V., 1997. Behaviour of different molecular-weight fractions of DOC of Elbe river water during river bank infiltration. Acta Hydrochimica Et Hydrobiologica 25, 145e150. Macleod, C.J.A., Semple, K.T., 2006. The influence of single and multiple applications of pyrene on the evolution of pyrene catabolism in soil. Environmental Pollution 139, 455e460.
S.B. Trinh et al. / Environmental Pollution 170 (2012) 95e101 Margoum, C., Malessard, C., Gouy, V., 2006. Investigation of various physicochemical and environmental parameter influence on pesticide sorption to ditch bed substratum by means of experimental design. Chemosphere 63, 1835e1841. Ministry of Agriculture Fisheries and Food, 2000. Biodegradation of Pesticides in the Unsaturated and Saturated Zones of Major UK Aquifers. Neal, C., Williams, R.J., Neal, M., Bhardwaj, L.C., Wickham, H., Harrow, M., Hill, L.K., 2000. The water quality of the River Thames at a rural site downstream of Oxford. Science of the Total Environment 251, 441e457. Pesce, S., Martin-Laurent, F., Rouard, N., Robin, A., Montuelle, B., 2009. Potential for microbial diuron mineralisation in a small wine-growing watershed: from treated plots to lotic receiver hydrosystem. Pest Management Science 65, 651e675. Pesce, S., Martin-Laurent, F., Rouard, N., Robin, A., Montuelle, B., 2010. Evidence for adaptation of riverine sediment microbial communities to diuron mineralization: incidence of runoff and soil erosion. Journal of Soils and Sediments 10, 698e707. Pieuchot, M., PerrinGanier, C., Portal, J.M., Schiavon, M., 1996. Study on the mineralization and degradation of isoproturon in three soils. Chemosphere 33, 467e478. Rae, J.E., Cooper, C.S., Parker, A., Peters, A., 1998. Pesticide sorption onto aquifer sediments. Journal of Geochemical Exploration 64, 263e276. Ray, C., Melin, G., Linsky, R.B., 2002a. Riverbank Filtration: Improving Source-water Quality. Kluwer Academic Publishers, The Netherlands. Ray, C., Soong, T.W., Lian, Y.Q., Roadcap, G.S., 2002b. Effect of flood-induced chemical load on filtrate quality at bank filtration sites. Journal of Hydrology 266, 235e258. Reid, B.J., Papanikolaou, N.D., Wilcox, R.K., 2005. IPU-undosed and induced isoproturon catabolic activity in dissimilar soils and soils under dissimilar land use. Environmental Pollution 133, 447e454. Scheunert, I., Reuter, S., 2000. Formation and release of residues of the C-14labelled herbicide isoproturon and its metabolites bound in model polymers and in soil. Environmental Pollution 108, 61e68.
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Sorensen, S.R., Aamand, J., 2001. Biodegradation of the phenylurea herbicide isoproturon and its metabolites in agricultural soils. Biodegradation 12, 69e77. Sorensen, S.R., Bending, G.D., Jacobsen, C.S., Walker, A., Aamand, J., 2003. Microbial degradation of isoproturon and related phenylurea herbicides in and below agricultural fields. Fems Microbiology Ecology 45, 1e11. Vargha, M., Takáts, Z., Márialigeti, K., 2005. Degradation of atrazine in a laboratory scale model system with Danube river sediment. Water Research 39, 1560e1568. Verstraeten, I.M., Thurman, E.M., Lindsey, M.E., Lee, E.C., Smith, R.D., 2002. Changes in concentrations of triazine and acetamide herbicides by bank filtration, ozonation, and chlorination in a public water supply. Journal of Hydrology 266, 190e208. Walker, A., Welch, S.J., 1992. Further-studies of the enhanced biodegradation of some soil-applied herbicides. Weed Research 32, 19e27. Walker, A., Jurado-Exposito, M., Bending, G.D., Smith, V.J.R., 2001. Spatial variability in the degradation rate of isoproturon in soil. Environmental Pollution 111, 407e415. Warren, N., Allan, I.J., Carter, J.E., House, W.A., Parker, A., 2003. Pesticides and other micro-organic contaminants in freshwater sedimentary environments e a review. Applied Geochemistry 18, 159e194. Weiss, W.J., Bouwer, E.J., O’Melia, C.R., Aboytes, R., Le, B.T., Schwab, K.J., 2004. Field and column studies to evaluate removal of pathogens and potential surrogate parameters during riverbank filtration. Abstracts of Papers of the American Chemical Society 228. U633eU633. Worrall, F., Parker, A., Rae, J.E., Johnson, A.C., 1996. Equilibrium adsorption of isoproturon on soil and pure clays. European Journal of Soil Science 47, 265e272. Younger, P.L., Mackay, R., Connorton, B.J., 1993. Streambed sediment as a barrier to groundwater pollution: insights from fieldwork and modelling in the river Thames basin. Journal of the Institution of Water and Environmental Management 7, 577e585.
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
Mechanistic insights into the role of river sediment in the attenuation of the herbicide isoproturon Son B. Trinh 1, Kevin M. Hiscock, Brian J. Reid* School of Environmental Sciences, University of East Anglia, Norwich NR4 7TJ, UK
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 December 2011 Received in revised form 14 May 2012 Accepted 21 May 2012
Mechanistic insights into the relative contribution of sorption and biodegradation on the removal of the herbicide isoproturon (IPU) are reported. 14C-radiorespirometry indicated very low levels of catabolic activity in IPU-undosed and IPU-dosed (0.1, 1, 100 mg L1) river water (RW) and groundwater (GW) (mineralisation: <2%). In contrast, levels of catabolic activity in IPU-undosed and IPU-dosed river sediment (RS) were significantly higher (mineralisation: 14.5e36.9%). Levels of IPU catabolic competence showed a positive log-linear relationship (r2 ¼ 0.768) with IPU concentration present. A threshold IPU concentration of between 0.1 mg L1 and 1 mg L1 was required to significantly (p < 0.05) increase levels of catabolic activity. Given the EU Drinking Water Directive limit for a single pesticide in drinking water of <0.1 mg L1 this result suggests that riverbed sediment infiltration is potentially an appropriate ‘natural’ means of improving water quality in terms of pesticide levels at concentrations that are in keeping with regulatory limits. Ó 2012 Elsevier Ltd. All rights reserved.
Keywords: Isoproturon Sorption Biodegradation Catabolic activity Sediment River water Groundwater Riverbank filtration
1. Introduction The abstraction of GW as a drinking water resource is commonplace throughout the world. Some of this drinking water relies upon bank infiltration to maintain its potability. To this end, undesirable impurities such as dissolved organic carbon (DOC) (Ludwig et al., 1997; Hoppe-Jones et al., 2010), pathogens (Weiss et al., 2004), and herbicides (Verstraeten et al., 2002; Vargha et al., 2005) are removed as water ingresses through the riverbank and is, in effect, ‘filtered’. The capacity of the natural system to fulfil this service is dependent upon a number of factors such as river sediment properties (Younger et al., 1993; Margoum et al., 2006), microbial activity (Warren et al., 2003), contaminant load (Helweg et al., 1998; Ray et al., 2002b), hydrogeological properties (Hiscock and Grischek, 2002) and integration of these factors. Abstraction of drinking water from boreholes adjacent to rivers has been practiced throughout Europe, the United States of America (USA) and in many areas of the developing world (Ray et al., 2002a). The herbicide isoproturon (IPU; 3-(4-isopropylphenyl)-1,1dimethyl urea) is of significant concern on account of its * Corresponding author. E-mail address: [email protected] (B.J. Reid). 1 Present address: Mekong River Commission, 184, Fa Ngum road, Unit18, Sithane Neua, Sikhottabong District, Vientiane 0100, Laos. 0269-7491/$ e see front matter Ó 2012 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.envpol.2012.05.026
particularly high usage historically. Monitoring by the Environment Agency of England and Wales in 2006 found that IPU was the commonest cause of breaches of the 0.1 mg L1 European Union limit in surface waters. In Great Britain, more than 2.0 106 kg of IPU were annually applied on more than 1.8 106 treated hectares from 1990 until 2007 (CSL, 2008). A number of studies have revealed the capacity of agricultural soils to respond to IPU input and become catabolically enhanced (Bending et al., 2001; El-Sebai et al., 2005; Reid et al., 2005; El Sebaï et al., 2011). There remains, however, the need to understand the role of RS in acting as a zone of attenuation in riverbank filtration (RBF) schemes and in the processing of diffuse contaminants in the hyporheic zone. The main aims of the present work were to assess the potential for natural attenuation of IPU in aquatic systems. Towards these ends biodegradation capacities in three compartments: RW, GW and RS were made. This research provides understanding of the capacity for RBF to accommodate IPU at a range of concentrations and, by extension, the potential for the application of RBF as a viable mitigation approach to remove herbicide residues from surface water. 2. Materials and methods Full details regarding the collection of environmental materials used in this study, their handling and information regarding their physical and chemical
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properties, along with details regarding reagents used, IPU quantification and statistical methods, are provided in supporting information. Provided below are descriptions of the experiments undertaken: i) a 14C-radiorespirometric study to assess the role of IPU concentration and incubation time on levels of IPU catabolic competence in RW, GW and RS; ii) a fixed-bed column recirculation experiment to establish IPU fate in the presence of the binary pairings of sterile/non-sterile RS and sterile/non-sterile RW, and; iii) the subsequent assessment of catabolic competence, with respect to IPU, in material removed from the fixed-bed columns (after 18 d recirculation). 2.1. Assessment of IPU catabolic competence in RW, GW and RS Assessment of catabolic activity was assessed by 14C-radiorespirometry wherein generation of 14CO2 was used as a tracer of 14C-IPU mineralisation. By monitoring the temporal production of 14CO2 in a test system, the catabolic activity of the microbial communities in RW, GW and RS was determined. The 14 C-radiorespirometry procedure was adapted from that described in Allan et al. (2006). In brief, materials under assessment were contained within respirometry flasks. Flasks were spiked with 14C-IPU (1 kBq delivered in 100 mL of methanol; IPU was uniformly 14C-labelled on its phenyl ring) resulting in a final concentration of 14C-IPU of 0.5 mg L1. A glass scintillation vial (7 mL) containing 1 M NaOH (1 mL) suspended from a crocodile clip was used to absorb 14CO2. This vial was periodically removed and replaced with a fresh vial of 1 M NaOH. Ultima Gold scintillation fluid (6 mL) was added to the removed vials. Following 24 h in darkness, 14C-activity in the removed vial was established by liquid scintillation counting (Canberra Packard Tri-Carb 2900TR). The respirometry assay was undertaken for 30 d from the time of 14C-IPU inoculation. To assess catabolic activity in RW and GW, an aliquot (30 mL, n ¼ 3) was transferred to a respirometry flask. Subsequently, standards of 12C-IPU were spiked into the flasks to give final concentrations of 0 (for IPU-undosed treatments), 0.1, 1.0 and 100 mg L1 (for IPU-dosed treatments). The resultant ‘treatments’ were incubated at room temperature with shaking for up to 30 d prior to the addition of 14C-IPU (to assess levels of catabolic activity). Following incubation times of 0 d, 5 d, 10 d and 30 d, 14C-IPU was spiked to a complete set of treatment flasks (n ¼ 3) to initiate the assessment of catabolic activity. These respirometry assays were undertaken for a subsequent 30 d from the point at which 14C-IPU was added. To assess catabolic activity in RS, portions of RS (10 g) were weighed into respirometry flasks containing sterile distilled water (130 mL, n ¼ 3). These flasks were firstly spiked with 12C-IPU standards (prepared in MiliQ water and delivered in 100 mL) to give concentrations of 0 (for IPU-undosed treatments) 0.1, 1 and 100 mg L1 (for IPU-dosed treatments). The resultant ‘treatments’ were incubated at room temperature (15e20 C) with shaking (as before) for 0 d, 5 d, 10 d and 30 d. After these incubation periods, a volume of 100 mL was removed (the flasks were ‘rested’ without shaking over-night before removing this aliquot). These removed aliquots were subsequently used to determine residual 12C-IPU concentrations using a solid phase extraction (SPE) method (see Supporting Information). The flasks and their residual contents (10 g RS and 30 mL MilliQ water) were then used to assess levels of catabolic activity. Towards these ends, the respirometry flasks were spiked with 14C-IPU and a respirometry assay undertaken for 30 d from the point of 14C-IPU inoculation. 2.2. Fixed-bed column recirculation experiment The fixed-bed column recirculation experiment was set-up as shown in Fig. S1. Four treatments (n ¼ 3) were established using the RW and RS samples in either sterile or unsterilised states: Treatment 1 with both RW and RS sterilised (autoclaved: 121 C for 30 min); Treatment 2 with both RW and RS not sterilized; Treatment 3 with only RS sterilised while RW was not sterilised, and; Treatment 4 with only RW sterilised while RS was not sterilised. After packing RW and RS into the system, an IPU stock solution was spiked in to the glass reservoirs to give a final concentration of 100 mg L1. RW samples were removed from the reservoir periodically over the course of 18 d. The samples were filtered (Millex-GP, 0.22 mm) and kept in a cold room (4 C, darkness) before determining IPU concentrations by HPLC (see Supporting information for analytical details). 2.3. Assessment of IPU catabolic competence in fixed-bed column materials Following 18 d of recirculation, RS in the columns representing Treatments 3 and 4 (these treatments contained, respectively, sterile RS with non-sterile RW and non-sterile RS with sterile RW) were assessed with respect to their levels of IPU catabolic competence using the 14C-radiorespirometry method. Samples (10 g) were removed under aseptic conditions from the columns and placed in respirometry flasks containing sterile distilled water (30 mL). Three flasks were prepared from each column (of which n ¼ 3). Flasks were subsequently spiked with 14 C-IPU (see above) and the 14C-radiorespirometry assay undertaken as described above for a total of 10 d.
3. Results and discussion 3.1. IPU catabolic competence in riverbank materials 3.1.1. IPU catabolic competence in RW and GW Biodegradation of IPU in surface water and GW have received little attention previously. The results presented here are significant in that they indicate the IPU-undosed catabolic activity with regards to IPU in both RW and GW to be very low (<2% IPU mineralisation) (Fig. 1). Additionally, no significant (p < 0.05) changes in levels of catabolic activity in RW and GW were observed following IPU addition (at concentrations of 0.1, 1 and 100 mg L1) and following incubation for up to 30 d (Fig. 1). While Johnson et al. (1998) reported very limited decomposition of IPU in GW. To the best of our knowledge catabolic competence with respect to IPU in RW has not been reported to date.
A
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Incubation time (d) Fig. 1. Mineralisation extents following addition of 14C-IPU and 30 d of subsequent assay time in RW (A), GW (B) and RS (C) in un-dosed controls (black) and IPU-dosed treatments with IPU concentrations 0.1 mg L1 (white), 1 mg L1 (hatched) and 100 mg L1 (cross hatched). Errors are shown as standard errors (n ¼ 3). Note y-axis scales differ.
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3.1.2. IPU catabolic competence in RS Levels of catabolic activity in un-dosed RS were markedly higher (14.5 1.6%) than those in RW and GW (<2% IPU mineralisation) (Figs. 1 and 2). Higher levels are probably a result of the presence of competent microorganisms. Fries et al. (2008) reported the number of microorganisms (per g) in RS to be 100e1000 times greater than that in RW (per mL). In addition to increased cell numbers, the
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opportunity for microbial attachment on to a solid matrix (Warren et al., 2003) may have assisted in the proliferation of microbial populations where RS was present (these populations typically being associated with surfaces in the environment, rather than living free in solution (Warren et al., 2003)). It is noted that IPU degradation in non-sterile GW has been reported to be negligible in the absence of a solid matrix (Johnson et al., 1998). Mineralisation was observed to develop in 3 stages (Fig. 2). The first stage represented a lag period (in this study, a lag period is defined as the time elapsed between 14C-IPU addition and mineralisation exceeding 5%). Following this lag period, a second stage of marked mineralisation was observed. Finally, mineralisation was observed to plateau. Several previous studies on the degradation of IPU in different agricultural soils have reported that the extent of IPU mineralisation in the treatments with previous exposure to IPU to be significantly higher than that in the treatments without previous IPU exposure (Walker and Welch, 1992; Cox et al., 1996; El-Sebai et al., 2005; Reid et al., 2005).
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Assay time (d) Fig. 2. IPU spike after treatments (6). Errors
catabolic activity in river sediment (RS) following addition of a 14C-IPU 0 d (A), 5 d (B), 10 d (C) and 30 d (D) in undosed (C) and IPU-dosed with IPU concentrations of 0.1 mg L1 (B), 1 mg L1 (,) and 100 mg L1 are shown as standard errors (n ¼ 3).
3.1.3. Extent and rate of IPU mineralisation in RS With the exception of the 30 d incubation, all of the 0.1 mg L1 IPU concentration treatments had significantly (p < 0.05) lower extents of mineralisation compared with the un-dosed control at corresponding incubation periods. In contrast, where IPU concentrations were elevated to 1 mg L1, extents of mineralisation were observed to be significantly higher when compared to the control values at all incubation times (Figs. 1 and 2). Thus, catabolic activity was significantly (p < 0.05) enhanced following IPU addition but subsequently diminished with increasing incubation time. This change is attributed to temporal changes in IPU concentration (see below). Where IPU concentrations were elevated still further, to 100 mg L1, extents of mineralisation were again observed to be significantly (p < 0.05) higher when compared to the control values at all incubation times (Figs. 1 and 2). Furthermore, extents of mineralisation in the 100 mg L1 IPU-dosed treatments were also significantly greater than in the 1 mg L1 IPUdosed treatments at all corresponding incubation times (Figs. 1 and 2). Again, as was the case in the 1 and 100 mg L1 IPU-dosed treatments, the catabolic activity initially increased and then decreased in response to temporal changes in substrate level (see below). Maximum 14C-IPU mineralisation (36.9 2.7%) was observed in the 100 mg L1 IPU-dosed treatments after 5 d. It should be borne in mind that some of the IPU associated 14C will be incorporated into cellular biomass and as a consequence complete (100%) mineralisation is unlikely to be achieved (Macleod and Semple, 2006). Maximum rates of mineralisation were observed in the 100 mg L1 IPU-dosed treatments following 5 d incubation (4.74% d1) (Fig. S2). Comparable maximum rates of mineralisation in IPU-exposed agricultural soil (5.04% d1) have also been reported (Reid et al., 2005). Correlation between extent of mineralisation and the maximum rate of mineralisation revealed a strong linear relationship between these parameters (r2 ¼ 0.81; Fig. 3). This relationship suggests that both the extent of mineralisation and rate of mineralisation are equally valid proxies with which to define catabolic competence. To our knowledge, the reporting of these results is the first instance of IPU mineralisation potential being established for RS. Mineralisation of another phenylurea herbicide, diuron, was recently reported in RS microcosms (Pesce et al., 2010). The extent of diuron mineralisation was however found to be less than 5% after 30 d (varying from 5 to 10 % after 84 d) in RS not exposed to diuron, but, relatively high, up to 50% after 84 d, in the diuron-exposed treatments (Pesce et al., 2010). Results presented here and those
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6
Mineralization rate (% 14CO2 d-1) Fig. 3. Relationship between extent of mineralisation and maximum rate of mineralisation in IPU un-dosed controls (C) and IPU dosed treatments at concentrations of 0.1 (B), 1 (,) and 100 mg L1 (6). Errors are shown as standard errors (n ¼ 3).
of Pesce et al. (2010) provide consistent evidence of microbial catabolic competence to degrade phenylurea herbicides present in RS; and, that this catabolic ability can be enhanced in the presence of these herbicides. Comparison of these results for RS with laboratory microcosm experiments with agricultural soils and aquifer materials provides further context. Several authors have reported 14 C-IPU mineralisation, typically of the order of 5e25%, over assay times of 60e90 d at about 20 C in un-dosed soils (Kubiak et al., 1995; Lehr et al., 1996; Pieuchot et al., 1996; Larsen et al., 2000; Scheunert and Reuter, 2000) with 30e50% 14C-IPU being mineralised within 30 d at 15e20 C being reported for pre-exposed soils (Bending et al., 2001; Sorensen and Aamand, 2001; Sorensen et al., 2003). Similarly, Pesce et al. (2009) reported extensive mineralisation of diuron: up to 35.6% (after 109 days) in a vineyard site previously treated with diuron and >40% (after 21 days) in river sediment samples collected downstream of the vineyard site. Collectively, the evidence presented here (for IPU) and that reported by Pesce et al. (2010) (for diuron) indicated catabolic competence for phenylurea herbicides to be comparable in RS and agricultural soil. 3.1.4. Influence of residual 12C-IPU concentrations on IPU catabolic competence in RS The residual concentrations of IPU in dosed RS treatments decreased over the 30 d incubation period at all treatment concentrations. In the 0.1 mg L1 IPU treatment concentrations decreased as follows: 0.04 0.01 mg L1 (0 d), 0.03 0.01 mg L1 (5 d), 0.03 0.01 mg L1 (10 d), and below detection limit (30 d) (the detection limit was 1 mg L1, however, samples were concentrated by 100 times by SPE before injection on to the HPLC). IPU residual concentration also decreased to below detection limit in the 1 mg L1 treatments; changing as follows: 0.92 0.01 (0 d), 0.61 0.07 (5 d), 0.48 0.05 (10 d) mg L1, and below detection limit (30 d). Although detectable IPU concentrations remained in the 100 mg L1 treatments after 30 d of incubation significant (p < 0.05) successive decreases in IPU concentrations were observed over the incubation period; changing as follows: 86.20 0.86 mg L1 (0 d), 76.31 2.10 mg L1 (5 d), 57.79 2.72 mg L1 (10 d) and 31.31 4.42 mg L1 (30 d). Remaining concentration of IPU presented in the treatment incubations at the time of 14C-IPU addition (i.e. the point at which catabolic competence assays were initiated) was observed to have a strong (r2 ¼ 0.768) log-linear relationship with levels of catabolic activity (Fig. 4). This evidence supports the perspective that levels
10
0 10
10
10
10
10 -1
Solution phase IPU concentration (µg L ) Fig. 4. Relationship between the extent of IPU mineralisation and the concentrations of IPU present in the IPU-dosed treatments at the time of 14C-IPU spiking in the 0.1 (B), 1 (,), and 100 mg L1 (6) treatments. Errors are shown as standard errors (n ¼ 3).
of catabolic activity are proportional to concentrations of IPU present. Collectively, the results reported (in Figs. 1, 2 and 4) indicate that a threshold concentration of IPU between 0.1 mg L1 and 1 mg L1 was required to promote a significant increase in catabolic competence. Significantly, the EU Drinking Water Directive (CEC, 1998) limits concentrations of a single pesticide in drinking water to < 0.1 mg L1. Thus, the results reported here suggest that should this drinking water limit be exceeded by a factor of about ten, then the RWeRS system should have the capacity to become catabolically competent and reduce IPU concentrations. In this way we suggest that the RWeRS system is potentially an appropriate means of naturally improving water quality in terms of pesticide levels at concentrations that are in keeping with regulatory limits. It is acknowledged that this conclusion is based upon the results reported here and that these results are inextricably linked to the specific sediment sampled and its exposure history to IPU. It is well recognised that sediment texture and its organic mater content exert considerable control over organic compound partitioning (Chung and Alexander, 1998, 1999). This partitioning, in turn, governing free herbicide concentrations and their influence upon substrate-dependent catabolic activity. The sediment used in this research was of high sandy content (79% sand/20% silt (Table S2)) and of low organic matter content (0.8% O.C. (Table S1)). Thus IPU present would have been of relatively high bioavailability in the treatments. Regarding pre-exposure to IPU, Neal et al. (2000) reported concentrations of several pesticides, including the phenylurea herbicides IPU and diuron, in the Thames River at an areas upstream of the sampling site. IPU concentrations were reported to increase at higher flows while diuron concentrations showed no association with flow rate. The average concentration of IPU in river water was 0.16 mg L1 while the maximum IPU concentration was 1.63 mg L1 (Neal et al., 2000). It is noted that this average value is close to that at which no enhancement in catabolic activity was observed in the respirometer assays (Fig. 2) while the maximum concentration is in excess of the concentration required to significantly increase catabolic activity in RS (Fig. 2). 3.2. Sorption and biodegradation of IPU in a RWeRS system Attenuation of IPU was divided into three phases (Fig. 5): Phase I, a ‘sorption phase’ occurred during the first day (consistent for all treatment types); Phase II, a lag or ‘adaptation phase’ (observed in
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100
-1 IPU concentration in RW (µg L )
90 80 70 60 Treatment 1 - Sterile RW and RS Treatment 2 - Both non-sterile Treatment 3 - Sterile RS (non-sterile RW) Treatment 4 - Non-sterile RS (sterile RW)
50 40 30 20 10 0 0
2
4
6
8
10
12
14
16
18
Recirculation time (d)
Fig. 5. Attenuation of IPU in a RWeRS system with sterile and non-sterile RW and RS, error bars present standard error of three replicates. RW: river water; RS: river sediment.
Treatments 2 and 4 only; these treatments both containing nonsterile RS) with ‘adaptation’ taking place between days 2 and 6, and; Phase III a ‘degradation phase’, (again only observed in Treatments 2 and 4) that took place between days 7 and 12. 3.2.1. Phase I e sorption phase Concentration of IPU in RW was rapidly decreased following the first day in all treatments (Fig. 5): Treatment 1 (15.8 1.1% sorption), Treatment 2 (16.7 0.5% sorption), Treatment 3 (14.9 2.5% sorption) and Treatment 4 (17.8 2.2% sorption). IPU sorption in all four treatments was not significantly different (p > 0.05). This suggests that sterilization did not significantly affect the sorption capacity of the RS. The average IPU sorption across all four treatments was determined to be 16.3 0.7%. This result is consistent with the values 9e37 % IPU sorption onto ditch sediment (Margoum et al., 2006). In addition, a sorption phase was noted to occur during the first recirculation day of the experiment; again, this observation being consistent with the sorption equilibrium period of ‘less than one day’ reported for several herbicides (including IPU) onto ditch sediment (Margoum et al., 2006). IPU concentrations did not significantly decrease (p > 0.05) over the subsequent 5 d of recirculation in any of the treatments, with IPU concentrations in Treatments 1, 2, 3 and 4 having decreased by only a further: 4.7 0.9%, 4.7 2.0%, 2.6 1.0% and 3.3 1.6%, respectively (Fig. 5). Based upon the results observed in the sorption phase, several sorption isotherm parameters of IPU in the RWeRS system were calculated using the data for all four treatment types collectively (see Supporting information), these parameters were: the maximum sorption capacity of the RS (CS,max) ¼ 225 10 mg kg1; solid-water distribution coefficient (KD) ¼ 2.80 0.14 L kg1; organic carbon-normalized partition coefficient (KOC) ¼ 349 18 L kg1; and the retardation factor (RD) ¼ 7.9 0.36. Using a first-order model, the average first order sorption rate constant ksorp (h1) based upon all four treatments was established to be 8.6 103 0.6 103. To our best knowledge, there are no previous reports of sorption characteristics of IPU to RS using a fixed-bed column method. Consequently, the contextualisation of our results is made with respect to the sorption of IPU to soils and other environmentally relevant geosorbents. A considerable range in IPU KD values (L kg1) has been reported: in top soils (0e50 cm) varying from 2.73 to 4.39; in subsoil (50e70 cm) varying from 0.75 to 0.91, and; in chalk (2.1e9.9 m below the surface) varying from 0.02 to 0.46 (Johnson et al., 1998). It is noted that, in the case of sorption of hydrophobic herbicides, the solid-water distribution coefficient (KD) can be
99
normalised to the organic-carbon content; yielding the organiccarbon-normalised distribution coefficient (KOC) (Karickhoff, 1984; Rae et al., 1998). The KOC value of IPU in surface soil samples (from 10 to 90 cm below surface with total organic carbon of 0.8%) was reported to be 150 L kg1 (Worrall et al., 1996). In agricultural soils, KOC values of IPU varied from 112 to 138 L kg1 (Worrall et al., 1996; Cooke et al., 2004). In aquifer sediments (depths varying 12e19 m) KOC values varied from 19 to 278 L kg1 (Rae et al., 1998). In sediment from the Kishon River (Israel) KOC values of IPU were reported to be 60.12 and 125.60 L kg1 (Chefetz et al., 2004). In comparison to these studies the KOC values of IPU reported here for RS are higher (349 18 L kg1). In this study, the sorption of IPU on to RS was noted to be similar to that in topsoil and much higher than in subsoil, chalk, and sediment with low TOC content. 3.2.2. Phase II e adaptation phase The concentration of IPU remained unchanged in Treatment 1 (abiotic control in which both RW and RS were sterilised) following initial sorption in the first day of recirculation; there was no significant decrease (p > 0.05) between the IPU concentrations between days 1 and 18 (Fig. 5). Similarly, no significant (p > 0.05) decreases in IPU concentration was observed in Treatment 3 (consisting of sterile RS and non-sterile RW) between days 1 and 18 (Fig. 5). In contrast, IPU loss from Treatments 2 and 4 (both containing non-sterile RS) differed significantly (p < 0.05) from those in Treatments 1 and 3 (both of which contained sterile RS) (Fig. 5). These treatments revealed an ‘adaptation phase’ wherein no significant (p > 0.05) difference in IPU concentrations were observed between days 1 and 5 of recirculation. Thereafter, a ‘degradation phase’ was observed in which IPU loss was rapid and complete (Fig. 5). 3.2.3. Phase III e biodegradation phase From days 7e12, the IPU concentrations in Treatments 2 and 4 decreased from, respectively, 74.2 2.5 mg L1 and 79.4 1.68 mg L1 to below the HPLC limit of detection. Degradation rates of IPU in Treatments 2 and 4 were estimated using a zeroorder model during Phase III only (see Supporting information) and were determined to be, respectively, 13.62 0.17 and 17.72 1.43 mg L1 d1. Half-lives of IPU in Treatments 2 and 4 were, therefore, calculated to be 2.67 0.04 d and 2.47 0.26 d, respectively. Biodegradation rates of IPU in Treatments 2 and 4 were not significantly different (p > 0.05). This suggests that although results from Treatment 2 indicated IPU could be degraded by microorganisms living in both RW and RS environments, results from Treatment 4 confirmed that RS-borne microorganisms were primarily responsible for the degradation of IPU. This finding is consistent with the results from the previous experiment that indicated very low levels of catabolic activity in RW (<2%) but significantly high levels of catabolic activity in RS (>14.5%). This finding supports the conclusion that RS has the potential to strongly influence degradation while RW does not. Owing to the absence of RS studies with which to draw comparisons, comparison is made with other geosorbents. Reported IPU degradation in systems containing: i) a combination of GW and topsoil, and ii) a combination of GW and chalk (3.1e7.2 m below the surface) showed that in both cases IPU was degraded to below detection level within 7e14 d (Johnson et al., 1998). A variation in IPU degradation to 50% of application ranging over 6.5e30 d within samples taken from the same field has been reported (Walker et al., 2001). An IPU 50% degradation (t50) has been reported to vary from 14 d to longer than 65 d in fields previously dosed and undosed with IPU (Bending et al., 2006). The IPU biodegradation rates in a GW-chalk system has
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been reported to vary with half-lives from 111 to 273 d, while the half-lives of IPU in topsoil has been reported to be shorter, from 15 to 34 d (Ministry of Agriculture Fisheries and Food, 2000). Thus, the half-lives reported here for non-sterile RS (of the order of 2.5 d) indicate a considerable capacity for RS to remove IPU. 3.3. Enhancement of IPU catabolic competence in RS samples taken from the recirculation experiment RS removed from the fixed-bed columns following 18 d of recirculation showed marked differences in their levels of IPU catabolic competence (Fig. S3). RS removed from Treatment 3 (comprised of sterile RS and non-sterile RW) indicted low levels of catabolic competence; with the extent of 14C-IPU mineralisation reaching only 1.6 0.1% (after 10 d; Fig. S3). This low level of mineralisation suggests minimal competence of RW-borne microorganisms to degrade IPU. This result is consistent with the finding for Treatment 3 during the recirculation experiments wherein no further IPU loss was observed following the initial sorption phase (Fig. 5). In addition, this result is also consistent with the low IPU level of IPU-undosed catabolic competence in RW (<2% after 30 d; Fig. 1). In contrast, RS removed from the fixed-bed column of Treatment 4 (comprised of sterile RW and non-sterile RS) exhibited high levels of IPU catabolic competence. The maximum extent of mineralisation in samples removed from Treatment 4 was 40.0 0.8% (after 10 d; sufficient time for mineralisation to plateau) (Fig. S3). A maximum 14C-IPU mineralisation rate of 29.4% d1 was observed in the first day of the assay (Fig. S3). This high level of mineralisation and rapid mineralisation rate suggests significant competence of RW-borne microorganisms to completely degrade IPU. While the extent of IPU mineralisation observed in RS samples removed from the recirculation experiment (40.0 0.8%) was similar to that observed in the respirometer assays (reaching a maximum extent of 36.9 2.7% in the 100 mg L1 treatment after 5 d) the rate of mineralisation in samples obtained from the recirculation regime (29.4% d1) was far greater than that observed in the initial screening of IPU concentration influence upon catabolic activity (4.74% d1). This evidence supports the suggestion that the final biodegradation phase of the recirculation experiments was responsible for the rapid loss of IPU in Phase III. 4. Conclusions Based upon the results obtained from both the recirculation experiment and 14C-radiorespirometry assays, it is concluded that IPU degradation in RSeRW systems is most strongly dependent upon microbial competence in RS and not in RW. Thus, while opportunities for IPU degradation are low for RW and GW, opportunities are much higher in RS. Results indicate that a threshold IPU concentration of between 0.1 mg L1 and 1 mg L1 was required to promote significant increases in catabolic competence in this study. This value is consistent with the EU Drinking Water Directive (CEC, 1998) that limits concentrations of a single pesticide in drinking water to < 0.1 mg L1. Results reported here suggest that should this limit be exceeded, by a factor of about ten, then the RWeRS system should have the capacity to become catabolically competent and reduce IPU concentrations. In this way we suggest that the RWeRS system is potentially an appropriate means of naturally improving water quality in terms of pesticide levels at concentrations that are in keeping with regulatory limits. This finding is important in the context of RBF as it supports the use of RBF as an appropriate means of water purification where pesticides in RW are elevated above regulatory limits.
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