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Diffusive gradients in thin films for predicting methylmercury bioavailability in freshwaters after photodegradation
Chemosphere 131 (2015) 184–191
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Diffusive gradients in thin films for predicting methylmercury bioavailability in freshwaters after photodegradation C. Fernández-Gómez, J.M. Bayona, S. Díez ⇑ Environmental Chemistry Department, Institute of Environmental Assessment and Water Research, IDÆA-CSIC, E-08034 Barcelona, Spain
h i g h l i g h t s Assessment of DGT-labile MeHg during photodegradation was successfully tested. The action of sunlight seems not to alter the lability of MeHg. The kpd MeHg is mainly affected by DOM, which plays a key role in light attenuation. The quality of DOM, rather than the quantity, is key in the bioavailability of MeHg.
a r t i c l e
i n f o
Article history: Received 21 November 2014 Received in revised form 12 February 2015 Accepted 25 February 2015
Handling Editor: X. Cao Keywords: Methylmercury Photodegradation DGT Bioavailable Polyacrylamide
a b s t r a c t Determination of the dissolved-bioavailable fraction of methylmercury (MeHg) and its degradation pathways in freshwaters deserve attention, to further our understanding of the potential risk and toxicity of MeHg. Since the photodegradation of MeHg is the most important known abiotic process able to demethylate MeHg, this study investigated the role of sunlight on MeHg bioavailability in freshwater environments. Experiments to calculate photodegradation rate constants of MeHg in different types of freshwater in combination with experiments to distinguish the labile fraction of MeHg after being exposed to sunlight were performed. The ability of diffusive gradients in thin films based on polyacrylamide (P-DGT) to assess DGT-labile MeHg during photodegradation was successfully tested. First order photodegradation rate constants (kpd) of bioavailable MeHg determined in five different types of waters with different amount of dissolved organic matter (DOM), were in the range 0.073–0.254 h1, confirming previous findings that once there is DOM in solution, which would favour the photodegradation process, the kpd is mainly affected by light attenuation. Simulated sunlight seems not to alter the lability of MeHg, although photodegradation processes may decrease the concentrations of MeHg, contributing to reduce the amount of bioavailable MeHg (i.e. MeHg uptake by DGT). However, the quality of DOM, rather than the quantity, plays an important role in the bioavailability of MeHg in freshwater. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction Methylmercury (MeHg) is one of the most widespread waterborne contaminants and tends to bioaccumulate and biomagnify throughout aquatic trophic chains (Clayden et al., 2013; Lavoie et al., 2013). Therefore, determination of MeHg degradation rates in aqueous matrices and identification of the factors involved in its bioavailability (i.e., available for uptake by biota) are crucial for a better understanding of the potential risk and toxicity of mercury. It has been shown that MeHg can be demethylated in the water column of lakes by the action of microorganisms (Matilainen and Verta, 1995; Schaefer et al., 2004), but recently, ⇑ Corresponding author. Tel.: +34 93 4006100; fax: +34 93 2045904. E-mail address: [email protected] (S. Díez). http://dx.doi.org/10.1016/j.chemosphere.2015.02.060 0045-6535/Ó 2015 Elsevier Ltd. All rights reserved.
abiotic processes driven by sunlight have received increasing attention (Sellers et al., 1996; Lehnherr and Louis, 2009; Hammerschmidt and Fitzgerald, 2010; Black et al., 2012). The photodegradation of MeHg is the most important known abiotic process capable of demethylating MeHg and has been suggested to be the principal sink for MeHg in lakes (Sellers et al., 2001; Lehnherr and Louis, 2009) consuming up to 80% of MeHg total input (Hammerschmidt et al., 2006; Hines and Brezonik, 2007). Visible and, more efficiently, ultraviolet (UV) radiation (Lehnherr and Louis, 2009; Black et al., 2012) are able to decompose MeHg, mainly by indirect photolysis, involving reactive oxygen species (ROS, e.g., OH, O2H, and 1O2) (Suda et al., 1993; Gardfeldt et al., 2001; Chen et al., 2003; Zhang and Hsu-Kim, 2010a) formed by sunlight irradiation of water and its dissolved organic matter (DOM; i.e. chromophoric groups in humic substances, such as
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quinones) (Alegria et al., 1997; Grandbois et al., 2008) or dissolved redox active inorganic components such as nitrate (Zepp et al., 1987) and iron (Hammerschmidt and Fitzgerald, 2010). Apart from its role in the formation and protection of ROS, DOM has been shown to attenuate UV light (Morris et al., 1995; Williamson et al., 1996) and thereby the MeHg photolysis process. Other variables such as solar radiation intensity and spectral composition, water constituents other than DOM (e.g., salinity, nitrate, or photoreactive trace metals), have been suggested to play a role in the photodegradation of MeHg (Zepp et al., 1987; Sellers et al., 1996; Chen et al., 2003; Hammerschmidt and Fitzgerald, 2006, 2010; Lehnherr and Louis, 2009; Black et al., 2012). In freshwater environments, both inorganic mercury and MeHg can be partitioned between suspended particulate matter (SPM), dissolved and colloidal water phases. The last is associated with the DOM occurring in freshwater ecosystems (Babiarz et al., 2001; Hill et al., 2009). This influences the speciation and potentially the bioavailability of mercury. Furthermore, the effect of solar radiation on the equilibrium among the different Hg species and its role in their bioavailability are still unknown. Determining the bioavailability of mercury is also essential in order to assess its risk and potential effects on exposed biota. For this purpose, the technique of diffusive gradients in thin films (DGT) can be used. This technique is widely used for passive sampling of trace metals, and allows the determination of timeweighted average concentrations of bioavailable contaminant species (Cusnir et al., 2014). The DGT technique has been successfully used for in situ determination of kinetically labile metal species in aquatic systems, such as MeHg (Clarisse et al., 2012) and is an effective tool for the assessment of bioavailability of Hg species (Amirbahman et al., 2013). The principle of the DGT technique is based on the diffusion of the dissolved species through a diffusive layer and their accumulation in an ion-exchange resin. A hydrogel and a membrane filter are commonly used as the diffusive layer and the resin is incorporated into a polyacrylamide gel. These three layers are enclosed and sealed in a small plastic device, so that only the membrane is exposed to the deployment solution. The timeaveraged concentration of the metal in the solution, C, can be calculated according to the Fick’s first law of diffusion as:
C¼
M Dg DAt
ð1Þ
where D is the diffusion coefficient of the metal in the diffusive layer, t is the deployment time, A the exposure surface area, and Dg the thickness of the diffusive layer. The mass (M) of the analyte accumulated by the resin is experimentally measured and provides the average labile metal concentration during the exposure time. The objective of this work was to study the photodegradation of MeHg in natural waters and to assess whether this process affect the bioavailability of MeHg. To accomplish this, we combined the calculated experimental photodegradation rate constants (kpd) in different types of freshwater with the bioavailable fraction of MeHg assessed by the DGT technique. Differences in photodegradation rates among the different types of freshwater were evaluated in terms of DOM concentration and conductivity, whereas differences in MeHg bioavailability were explained based on DOM concentration and quality, electrical conductivity, and the equilibrium binding constants of the MeHg species present. 2. Materials and methods 2.1. DGT manufacturing An in-house manufactured polyacrylamide-based DGT (P-DGT) device was used for assessment of labile or bioavailable MeHg. Details of diffusive gel, resin gel preparation, and DGT device
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assembly have been outlined previously (Fernandez-Gomez et al., 2011, 2014). Briefly, the binding layer, also known as the resin gel, consisted of a 3-mercaptopropyl-functionalised silica gel embedded in a polyacrylamide gel (0.4 mm thick). On top of the resin gel was a 0.4 mm thick disk of polyacrylamide gel used as the diffusive gel that was covered by a 0.1 mm thick, 0.45 lm pore size Nylon filter membrane. The reagents and materials employed for the preparation of the DGT gels and for MeHg determination are described in previous works (Fernandez-Gomez et al., 2011, 2012). Thiourea (ACS, P99.0%), purchased from Fluka (Steinheim, Germany), and 37% fuming hydrochloric acid (ACS, ISO), purchased from Merck (Germany), were used to prepare an acid thiourea solution for MeHg elution from resin gels. Methylmercury chloride (CH3HgCl, 99%) was obtained from Strem (Newburgport, MA, USA), and dimethylmercury chloride ((CH3)2HgCl) from Alfa Aesar (Karlsruhe, Germany). Stock standard solutions were prepared at 1000 mg L1 (as Hg) in acetone and stored at 20 °C. Working solutions were prepared weekly by diluting the stock solutions with acetone to a range of 0.02–500 lg L1 (as Hg). 2.2. Apparatus Gas chromatography (GC) coupled to atomic fluorescence spectrometry via a pyrolytic reactor (GC-Py-AFS) was used for the analysis of MeHg in water samples and DGT resin gel eluates. The GC analysis was accomplished with a non-commercial system formed by a Thermo Trace GC ultra (Milan, Italy) gas chromatograph interfaced to an AFS Tekran Model 2500 (Toronto, Canada) detector via a pyrolyser (Hg-800, Rektorik R&D Chromatography, Meyrin, Switzerland). Details of GC analysis have been described previously (Carrasco et al., 2009). 2.3. Freshwater origin Five different types of water were used: MilliQ water (NaCl 0.01 M) (MQW); MilliQ water (NaCl 0.01 M) with organic matter (10 mg DOC L1) from a Nordic reservoir NOM (IHSS, 1R108N) (MQO); water from the Ebro River at Flix reservoir, Catalonia, Spain (FLX); and water taken at two sites along a freshwater wetland-lake gradient in Boreal Sweden. FLX water was collected at the Flix reservoir (41°140 N, 0°320 E) on the course of the Ebro River (annual mean = 255–424 m3 s1; watershed area 85 820 km2). The Flix reservoir is relatively small (area = 320 ha; volume = 11 hm3), with a very short water residence time of 0.15 d. The two Swedish sites were a dystrophic lake, Ängessjön (ANG; 64°20 5600 N 20°500 1700 E), and a mixed Sphagnum/Carex peatland-dystrophic lake with approximately 30% open water, Kroksjön (KSN; 63°570 800 N 20°380 1300 E). The ANG lake has an average water depth of about 2–2.5 m and more than 50 cm deep organic sediments. About 30% of the shores are covered by peatland and 70% by shallow mineral soils. The KSN site is dominated by floating mats and islands of Sphagnum/Carex peat mixed with small areas with open water of a maximum water depth of about 1 m. 2.4. Calibration of DGT units for MeHg measurements In order to measure MeHg in the five different waters described above, two MeHg solutions (1 lg L1 MeHgCl, 0.01 M NaCl), one with organic matter from a Nordic reservoir (IHSS, 1R108N) (10 mg L1) and the other one without, were prepared in 5 L amber glass bottles. The pH was adjusted to 7 using 1M NaOH or HCl. Afterwards, the solution was left to equilibrate overnight while being stirred with a magnetic stirring bar. Ten DGT devices of PDGT were submerged in each solution, whose temperature was
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controlled and kept at 25.5 ± 0.5 °C. Duplicates of DGT units were retrieved after 2, 4, 8, 14, and 24 h. A duplicate that was not deployed was used as the DGT blank (0 h). At each sampling interval, 5 mL of each MeHg solution were collected and acidified to 0.4% with HCl to monitor the MeHg concentration remaining in solution, and 10 mL were taken only from the DOM-containing solution to measure dissolved organic carbon (DOC). Both the DGT units and the water samples were stored in the fridge until analysis within one week. 2.5. MeHg photodegradation experiment All experiments were conducted in acid-washed 500 mL widemouth TeflonÒ fluorinated ethylene propylene (FEP) bottles (Thermo Scientific-Nalgene, Rochester, NY, USA). Two bottles were filled with every water type; one replicate was used for the photodegradation process and the other as dark control. Every bottle was filled to the top, to which was added 250 ng Hg (CH3HgCl in acetone). The bottle was shaken vigorously, wrapped with aluminium foil, and left to equilibrate overnight at room temperature. In the case of the Swedish waters, since they had very high DOC levels, they were diluted to roughly 10 mg L1 of DOC. The following morning, the bottle was unwrapped and placed in the SuntestÒ for 7.5 h. Before light exposure (0 h) and at different time points during the experiment (0.5, 1.5, 3.5, and 7.5 h), the pH, conductivity, and DOC were determined. Moreover, at these time points, 2 mL of water were taken, preserved from light in an aluminium foil-wrapped glass vial, and acidified to 0.4% with HCl. Then 20 lL of ethylmercury chloride (EtHgCl; 230 ng L1) were added as internal standard and the water was stored at 4 °C until analysis. At the end of the photodegradation experiment, the bottle was double wrapped with aluminium foil and kept in the fridge until the bioavailability experiment (Section 2.6). The same procedure was used for dark controls, except that the bottles were double wrapped with aluminium foil also when placed inside the SuntestÒ equipment and aliquots were taken only three times (0, 3.5, and 7.5 h). The photodegradation rate constants (kpd) were calculated from first-order decay kinetics as the slope of the negative linear relationship between the natural logarithm of the quotient between the concentration of MeHg at the sampling occasion and the initial concentration, ln([MeHg]/[MeHgi]), and the time of exposure to the xenon lamp (h). 2.6. DGT-labile MeHg experiment The bottles that were exposed to photodegradation and those used as dark controls were used for this experiment. Two P-DGT units were introduced into FEP bottles exposed (n = 5) and unexposed (n = 5) to sunlight. All bottles were double wrapped with aluminium foil and placed in the shaking incubator Innova 40R (New Brunswick Scientific, Edison, NJ, USA) under conditions of 25 °C and 30 rpm. After 23 h, the DGT units were retrieved and stored in the fridge (4 °C). To monitor the MeHg concentration in the water throughout the experiment, 2 mL aliquots of water were taken from each bottle before and after the DGT deployment. 2.7. Calculation of the photodegradation yield index The photodegradation yield index (PDY) was estimated by decoupling the global spectral irradiation in specific values. This calculated PDY is not an experimental value but it should be useful for estimating the relative amount of MeHg photodegraded in the middle of FEP bottles (i.e. the longitudinal axis of the cylindrical bottle). The overall MeHg PDY (PDYoverall) is the addition of the specific PDY at different wavelengths (UVA, UVB, and visible light) (Eq. (2)).
PDYoverall ¼ PDYVis þ PDYUVA þ PDYUVB
ð2Þ
For each wavelength, the PDY is calculated as the product of three fractions: downwelling irradiance in the middle of the bottle (fd), contribution in the photodegradation process (fp), and contribution in the spectrum of the light source (i.e. the lamp) used for the experiment (fs), as follows:
PDYVis ¼ fdVis fpVis fsVis PDYUVA ¼ fdUVA fpUVA fsUVA
ð3Þ
PDY UVB ¼ fdUVB fpUVB fsUVB The fd at different wavelengths (Vis, UVA and UVB) were calculated using the wavelength specific light attenuation coefficients (Kd) described by Morris et al. (1995) and the downwelling irradiance equation described by Williamson et al. (1996) (Eq. (4)), where z is the radius of the bottle.
fdVis ¼ exp½ðK dVis Þz fdUVA ¼ exp½ðK dUVA Þz
ð4Þ
fdUVB ¼ exp½ðK dUVB Þz The different fp were obtained from the wavelength specific first order photodegradation rate constants (kpd): 0.0023 ± 0.0002, 0.10 ± 0.024, and 7.2 ± 1.3 m2 E1, for kpdPAR, kpdUVA, and kpdUVB, respectively, resulting in a relative ratio of 1:43:3100 for the three constants, reported in a previous study (Fernandez-Gomez et al., 2013) (Eq. (5)).
fpVis ¼
1 ; 3144
fpUVA ¼
43 ; 3144
fpUVB ¼
3100 3144
ð5Þ
Finally, the fs was obtained from the global spectral distribution provided by the instrument supplier of the xenon lamp used in the experiment (50%, 5.6%, and 0.4% of visible light, UVA, and UVB, respectively) (Eq. (6))
fsVis ¼ 0:50;
fsUVA ¼ 0:056;
fsUVB ¼ 0:004
ð6Þ
2.8. Analytical methods 2.8.1. Elution of MeHg from DGT resin gels After being retrieved from the DGT assembly, every resin gel was placed in a glass vial using hydrochloric acid (10%)-cleaned tweezers. Two millilitres of a freshly prepared thiourea acid solution (1.3 mM thiourea, 0.1 M HCl) were added to each vial, which was wrapped with aluminium foil to prevent photodegradation, and left at room temperature for the elution to take place. Following 24 h of exposure, the vials were stored in the fridge (4 °C) until analysis (within a week). 2.8.2. MeHg determination A 250 lL aliquot of resin extract was placed in a 6 mL glass vial with a polytetrafluoroethylene (PTFE)-coated silicone rubber septum containing 3 mL of citric-citrate buffer (pH = 4.5–5). Ten microlitres of EtHgCl (230 pg mL1 as Hg) in acetone were also added to the vial as instrumental standard. Once the vial was covered, it was placed in a water bath at 70 °C and stirred at 1200 rpm using an ETS-D4 fuzzy thermometer and a digital RCT hot plate, both purchased from Ika Labortechnik (Staufen, Germany). Then, 100 lL of 1% aqueous solution of NaBPh4 were added to the mixture through the septum as a derivatising agent. After a 5-min derivatisation step, extraction was conduted for 27 min by headspace solid-phase microextraction (HS-SPME) using a 100 lm polydimethylsiloxane (PDMS) fibre. The final detection was carried out using a GC-Py-AFS. Three resin gels were spiked with 3 ng of MeHg in order to calculate the recovery of the method, which was 65% (RSD = 4%, n = 3). This value was later used to correct
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the results. In the case of the water samples, the methodology was basically the same but with two differences: the EtHgCl was added as internal standard right after taking the samples and before storing them to correct for any possible loss of MeHg during storage or due to adsorption on the vial walls, apart from detector variations; and the aliquot of the sample for HS-SPME was 1 mL. Every time the instrument was switched on to run a set of samples, the repeatability of the methodology was tested by analysing the same sample, which yielded an RSD of less than 5% (n = 3), proving the robustness of the analytical procedure. A six-point calibration curve was performed in the range from 29 to 1180 pg of MeHg (as Hg). The limit of detection (LOD) was calculated as the mean of the method blanks plus three times their standard deviation, and the limit of quantification (LOQ) as the mean of the method blanks plus 10 times their standard deviation, resulting in 5.8 and 19 pg, respectively. 2.8.3. Statistical analysis The difference between the DGT-labile MeHg concentrations in the water that underwent photodegradation and in the dark control was analysed with a Mann–Whitney U test. Comparison of the diffusion coefficients of MeHg with and without DOM was performed by a two-sample t-test. The significance level was set at p 6 0.05. Statistical analysis was performed using the SPSS 15.0 for Windows software (SPSS Inc., Chicago, IL, USA). 3. Results and discussion 3.1. Calibration of DGT units for MeHg measurements The results of the calibration showed that DGT based on polyacrylamide accumulated MeHg in proportion to exposure time both in the presence and the absence of DOM (Fig. 1). The diffusion coefficient of MeHg in the diffusive layer can be calculated from the slope (s) of the relationship between the mass of MeHg accumulated by the DGT units and the deployment time. The diffusion coefficients of MeHg in the diffusive layer in the absence and in the presence of DOM in P-DGT were significantly different, with values of 2.54 106 and 1.67 106 cm2 s1, respectively, at 26 °C. These values agree with those that had been obtained previously (Fernandez-Gomez et al., 2014). Values of diffusivity for MeHg in the DGT diffusive layer in the absence and presence of DOM were not so different from each other as for inorganic Hg determined in a previous work (Fernandez-Gomez et al., 2011) which might suggest that MeHg-DOM complexes are more labile than Hg(II)-DOM complexes and have a higher dissociation rate
Fig. 1. Time-series experiment. Mass of MeHg ([MeHg]i = 1 lg L1, 0.01 M NaCl) accumulated in the resin at different deployment times, in the absence (MilliQ) and in the presence of DOM (10 mg L1).
187
constant. Moreover, this result highlights the importance of calibrating the DGT units in a solution with similar characteristics (e.g., ion strength, DOM, etc.) to those where DGT would be deployed. 3.2. MeHg photodegradation experiment As shown in Fig. 2, for all five types of water, the pH remained fairly stable during the experiment. However, the DOC concentration decreased slightly due to DOM photolysis (Mopper et al., 1991; Hongve, 1994). In Fig. 3, it can be observed that for all the waters the MeHg concentration diminished when the bottles were exposed to the xenon lamp. On the contrary, the dark controls presented constant MeHg levels throughout the experiment. The five aqueous matrices used in this experiment covered a wide range of pH (5.2–8.3). However, as reported previously pH does not significantly affect the kpd of MeHg (Sellers et al., 1996; Chen et al., 2003). Concerning the electrical conductivity, previous studies (Zepp et al., 1987; Chen et al., 2003) observed an enhancement in kpd of MeHg with increasing chloride concentration, and proposed that it was due to the aqueous oxidation of chloride by hydroxyl radicals, producing chlorine radicals that may attack the Hg–C bond. Taking these findings into consideration, the higher kpd in FLX water (Table 1) might be explained by its high electrical conductivity (1010 lS cm1) in comparison with the Swedish waters (ANG and KSN, 11–40 lS cm1). However, MQO and MQW waters, despite also having high conductivity (1150 lS cm1), presented lower MeHg kpd rate constants (0.106 and 0.073 h1, respectively). The reason for the poorer linearity (R2 = 0.89) of the kpd rate in MQW in contrast to the rest of the water types (Fig. 3) is likely the lack of organic matter, which has been shown to contribute to MeHg photodecomposition by means of oxygen reactive intermediates (OH and 1O2) (Alegria et al., 1997; Chen et al., 2003; Zhang and Hsu-Kim, 2010b) generated by the interaction of sunlight with DOM. Unlike previous works,
Fig. 2. Changes in DOC and pH during the course of the MeHg photodegradation experiment.
C. Fernández-Gómez et al. / Chemosphere 131 (2015) 184–191
(a)
0.2 0 -0.2
PD
-0.4
Dark y = -0.0729x -0.0143 R² = 0.8871
-0.6 -0.8
0
2
4
6
ln ([MeHg]/[MeHgi])
ln ([MeHg]/[MeHgi])
188
8
(b)
0.2 0 -0.2
PD
-0.4
Dark
-0.6
y = -0.1057x + 0.0466 R² = 0.9715
-0.8 -1
0
2
(c)
0 -0.4 PD
-0.8
Dark
-1.2 y = -0.2538x -0.0812 R² = 0.9903
-1.6 -2
0
2
4
6
8
ln ([MeHg]/[MeHgi])
ln ([MeHg]/[MeHgi])
0.4
0.2
6
8
(d)
0 -0.2 PD
-0.4
Dark
-0.6 y = -0.1178x + 0.0447 R² = 0.9881
-0.8 -1
0
Time (h)
ln ([MeHg]/[MeHgi])
4
Time (h)
Time (h)
2
4
6
8
Time (h)
(e)
0.2 0 -0.2
PD
-0.4
Dark
-0.6
y = -0.1119x + 0.0689 R² = 0.9734
-0.8 -1
0
2
4
6
8
Time (h) Fig. 3. Relationships between ln([MeHg]/[MeHgi]) and light exposure time (h). The kpd full spectrum (m2 E1) was determined as the negative of the slope of the linear relationship. (a) MilliQ water (NaCl 0.01 M) (MQW), (b) MilliQ water (NaCl 0.01 M) with organic matter (MQO), (c) water from the Ebro River (FLX), (d) water from Ängessjön lake (ANG), and water from, (e) Kroksjön peatland-lake (KSN). PD: photodegradation experiment; Dark: dark controls.
Table 1 Water quality parameters, lability and bioavailability information. Initial DOC concentration, average pH along the experiment, conductivity, MeHg photodegradation rate constant (kpd), ratios between the MeHg mass accumulated by one DGT unit (M) and the average MeHg concentration (measured directly) in the deployment water during sampling (C), and percentages of bioavailable MeHg in the dark control and in the water exposed to artificial sunlight. Water type
[DOC] (mg L1)
pH
kpd (h1)
Cond.(lS cm1)
M/C (Dark) (cm3)
M/C (Light) (cm3)
% Bioav. (Dark)
% Bioav. (Light)
ANG KSN FLX MQO MQW
10.0 12.3 2.1 9.3 60.010
6.7 5.7 8.3 5.4 5.2
0.118 ± 0.009 0.112 ± 0.013 0.254 ± 0.014 0.106 ± 0.013 0.073 ± 0.018
40 11 1010 1150 1150
4.6 ± 0.4 4.4 ± 0.4 6.7 ± 1.0 8.3 ± 1.5 6.9 ± 1.8
4.8 ± 0.8 5.1 ± 0.2 6.1 ± 0.5 7.8 ± 2.1 7.8 ± 1.0
52 ± 5 50 ± 5 76 ± 11 95 ± 17 53 ± 14
55 ± 9 58 ± 3 70 ± 5 88 ± 24 60 ± 8
which suggested that photodegradation of MeHg hardly occurs in the absence of DOM (Chen et al., 2003; Zhang and Hsu-Kim, 2010a) it has been recently reported that MeHg can be degraded by sunlight when it is dissolved in plain ultra-pure water in the presence of chlorides (Sun et al., 2013). In the same manner, we observed slight photodecomposition of MeHg even in our MilliQ water with NaCl 0.01 M in solution (Fig. 3a, kpd-MQW = 0.073 h1). The lower kpd found for MQO and boreal waters (i.e., AGN and KSN) in comparison with FLX could be attributed to a higher DOM concentration, which attenuated more intensely the light penetrating the bottle. This might confirm previous findings (Fernandez-Gomez et al., 2013) that unless the MeHg degradation reaction is not limited by the generation of free radicals or by the complexation of MeHg by thiol groups, the main effect of DOM on MeHg photodegradation seems to be light attenuation. 3.3. Photodegradation yield index In order to better explain the difference in the kpd values between FLX and DOM-dominated waters, the PDY index was
calculated. Eqs. (3)–(6) were employed to calculate fp, fs, and fd values, and subsequently, PDY of MeHg for every light spectral range (UVB, UVA and visible light) as it is shown in Table 2. Then, it was possible to compare two waters by matching the results obtained in the Eq. (2), e.g. PDYoverall. For example, the obtained PDY values for MQO and FLX waters were 0.0021 and 0.0041, respectively (Table 2). Thus, the PDYMQO:PDYFLX ratio was 1:2.0. This result can provide the most likely explanation of the difference between the kpd of those two waters, since the kpd-MQO:kpdFLX ratio was 1:2.4, quite similar to the PDYMQO:PDYFLX ratio. Therefore, these results suggest that the main variable affecting MeHg photodegradation might be the DOC concentration, which plays an important role in attenuation of the different light wavelengths. Obviously, there were other variables that were not analysed that could affect the photodecomposition process. Analogous results were obtained among the rest of our waters (PDYANG:PDYFLX = 2.1 vs. kpd-ANG:kpd-FLX = 2.2; and PDYKSN:PDYFLX = 2.7 vs. kpd-KSN:kpd-FLX = 2.3). Concerning the statistical comparison (t-student, a = 0.05) of the obtained kpd values, no significant difference was found, except
C. Fernández-Gómez et al. / Chemosphere 131 (2015) 184–191 Table 2 Calculation of MeHg photodegradation yield index (MeHg PDY). Fraction of contribution in the photodegradation process (fp), fraction of contribution in the spectral composition of the lamp (fs), fraction of downwelling irradiance reaching the middle of the bottle with water (fd), and PDY of MeHg for every light spectral range (UVB, UVA and visible light). UVB
UVA
Vis
fp fs
0.986 0.004
0.014 0.056
0.0003 0.500
fd FLX (2.1 mg OC L1) MQO (9.3 mg OC L1) ANG (9.9 mg OC L1) KSN (12.3 mg OC L1)
0.825 0.362 0.336 0.249
0.924 0.643 0.622 0.543
0.982 0.921 0.917 0.897
FLX MQO ANG KSN
PDYUVB
PDYUVA
PDYVIS
MeHg PDYoverall
0.0033 0.0014 0.0013 0.0010
0.0007 0.0005 0.0005 0.0004
0.0002 0.0001 0.0001 0.0001
0.0041 0.0021 0.0019 0.0015
in the case of FLX versus the rest of the waters (p 6 0.0008), and in the case of ANG versus SKM (p = 0.02). 3.4. Lability of MeHg in the different types of water As mentioned before, to accurately determine the labile fraction of Hg in water by DGT, a diffusion coefficient (D, Eq. (1)) obtained under similar conditions to those in the water in which the DGT units were deployed should be used. However, in this case we studied five different types of water with different characteristics, which made it difficult to have a DGT calibration (i.e., D determination) for each of them. Thus, to rigorously compare the lability of MeHg in those waters, we worked with the ratio between the MeHg mass accumulated by one DGT unit (M) and the average MeHg concentration (measured directly) in the tested water during sampling (C) (Table 1). In this manner, we did not have to deal with problems derived from the use of the same D for all the different waters. The results were subjected to a Mann–Whitney test to statistically compare the MeHg lability indicator (M/C) in the water that underwent photodegradation and in the dark control. The test showed no significant differences between the two treatments in any of the five types of water. Therefore, we can conclude that the action of sunlight seems not to change the lability of MeHg. Aside from the influence of solar radiation on MeHg lability, the differences among the five waters can be discussed. The high values of the equilibrium binding constants (log K) for the association of MeHg with thiol-containing compounds or humic substances (log KMeHg-DOM = 12.6–16.9) (Reid and Rabenstein, 1981; Hintelmann et al., 1997; Amirbahman et al., 2002; Qian et al., 2002) in comparison with the equilibrium binding constant of MeHgCl (log KMeHgCl = 5.25), at similar concentrations to those studied here, suggest that MeHg associates primarily with the thiol groups in the DOM. Thus, in freshwater systems, when DOM concentration is sufficiently high, the MeHg-DOM complex should be more abundant than the MeHgCl complex. Black et al. estimated that DOM bound >97% of the MeHg at a salinity of 25‰ and DOM concentration of 1.5 mg OC L1 when MeHg concentration was 0.25–1.25 ng L1. Therefore, under our experimental conditions MeHg-DOM binding should also be predominant. However, changes in the complexation of MeHg by DOM and chloride should be a function of DOM concentration, salinity, and pH. Thus, the lowest M/C ratios found in ANG (4.6–4.8) and KSN (4.4–5.1) waters (Table 1) could be explained by their high DOM concentration
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(P10 mg DOC L1) and their low conductivity (40 and 11 lS cm1, respectively), and hence by the prevalence of MeHgDOM complexes in solution. These complexes are bigger in size than MeHg+, which makes them diffuse at a slow rate through the DGT diffusive layer, and they dissociate with more difficulty within the diffusive gel. Consequently, since a smaller amount of MeHg is bound to the thiol groups in the resin gel, a minor proportion of MeHg will be measured. A similar phenomenon happens also when working with living organisms. Some authors stated that an increase in DOM concentration led to a decrease in MeHg bioavailability, as tested in algae, seston and hydropsychids, or bacteria (Gorski et al., 2008; Tsui and Finlay, 2011; Luengen et al., 2012; Ndu et al., 2012) as the MeHg-DOM complexes are too large or hydrophilic to pass through the cell membrane. Furthermore, the slightly higher M/C ratio in FLX water (6.1–6.7) could be accounted for by the same behaviour. This water had a lower DOM concentration (2 mg DOC L1) and a notably higher conductivity (1010 lS cm1) than the others. Therefore, there could be a higher amount of MeHgCl, which is more easily dissociable than MeHg-DOM complexes. Concerning MQO and MQW waters, the difference between them is that MQO had a DOM concentration of 9.3 mg L1 and MQW was DOM-free. Although log K was substantially higher for MeHg associated with humic substances than for MeHgCl, MQO presented a similar MeHg lability (7.8–8.3) as MQW (6.9–7.8). Again, the presence of organic matter seemed not to affect MeHg lability significantly, as shown by the similar values of diffusivity for MeHg in P-DGT in the absence and presence of DOM. On the other hand, in order to estimate the fraction of MeHg that could be available to biota in all the water types, the following approach was taken. The percentage of ‘bioavailable MeHg’ (Table 1) was calculated as the quotient between the DGT-labile MeHg (measured with DGT) and the average of the MeHg concentration (measured directly in the water before and after DGT deployment). The DGT-labile MeHg concentration was calculated using D of MeHg in the presence of DOM for all the waters except for MQW, for which D in the absence of DOM was employed. As stated before, the D used for the DGT calculations should be obtained from a calibration performed in water with similar or identical characteristics to those of the water in which the DGT units were deployed, if an accurate labile (and then, proxy for bioavailable) MeHg concentration determination is desired. The high percentage of bioavailable MeHg in MQO (88–95%) versus that in MQW (53–60%) does not match our previous claim, nevertheless it might be explained by Pickhardt and Fisher’s (Pickhardt and Fisher, 2007) findings, that showed that DOC may enhance uptake. These authors observed that MeHg bioaccumulation in eukaryotic cells was greater in high DOC water. This uptake enhancement could occur through association of DOM with cell surfaces, which may facilitate uptake of Hg species into cells and/or promote transport by influencing the permeability of the cell membrane (Vigneault et al., 2000; Boullemant et al., 2004). A similar phenomenon might occur in the DGT membrane filter and diffusion gel.
4. Conclusions Our findings show that the DGT technique was successful to monitor MeHg photodegradation. Values of diffusivity for MeHg in the P-DGT in the absence and presence of DOM suggest that MeHg-DOM complexes are more labile than Hg(II)-DOM complexes. Concerning the photodegradation study, it can be concluded that as long as there is organic matter in the solution, which favours the photodegradation process, the kpd of MeHg is mainly affected by the light attenuation effect by DOM. A new
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and simple index has been developed and successfully employed to calculate a rough and relative amount of MeHg photodegraded in the middle of the bottle, the so-called, photodegradation yield index, PDY. This index is very useful for comparing two waters by matching this index with the ratio of its kpd and for confirming that DOM is key in the photodegradation rate for MeHg in DOM (humic)-dominated waters. On the other hand, the action of sunlight seems not to change the lability of MeHg and, therefore, its bioavailability; however, the DOM quality could be the main variable affecting it. Moreover, photodegradation resulted in decreased concentrations of MeHg, and thus, should decrease the amount of MeHg that is bioavailable, or in more precise wording, taken up by the DGT sampler. Future works should be focused on studying the influence of the type of DOM or humic substances present in the water on the bioavailability of MeHg, since several studies have also suggested that DOM composition may determine site-specific bioavailability of MeHg. Acknowledgments Financial support was provided by the Spanish Ministry of Economy and Competitiveness, MINECO (Project CTQ201125614). CFG would like to express her gratitude to Spanish National Council of Research (CSIC) and European Social Fund (ESF) for a JAE-predoc fellowship. Special thanks are also due to Ulf Skyllberg’s group, for supplying some water samples. We acknowledge the comments of anonymous referees, which helped us to improve the manuscript. References Alegria, A.E., Ferrer, A., Sepulveda, E., 1997. Photochemistry of water-soluble quinones. Production of a water-derived spin adduct. Photochem. Photobiol. 66, 436–442. Amirbahman, A., Massey, D.I., Lotufo, G., Steenhaut, N., Brown, L.E., Biedenbach, J.M., Magar, V.S., 2013. Assessment of mercury bioavailability to benthic macroinvertebrates using diffusive gradients in thin films (DGT). Environ. Sci. Proc. Impacts 15, 2104–2114. Amirbahman, A., Reid, A.L., Haines, T.A., Kahl, J.S., Arnold, C., 2002. Association of methylmercury with dissolved humic acids. Environ. Sci. Technol. 36, 690–695. Babiarz, C.L., Hurley, J.P., Hoffmann, S.R., Andren, A.W., Shafer, M.M., Armstrong, D.E., 2001. Partitioning of total mercury and methylmercury to the colloidal phase in freshwaters. Environ. Sci. Technol. 35, 4773–4782. Black, F.J., Poulin, B.A., Flegal, A.R., 2012. Factors controlling the abiotic photodegradation of monomethylmercury in surface waters. Geochim. Cosmochim. Acta 84, 492–507. Boullemant, A., Vigneault, B., Fortin, C., Campbell, P.G.C., 2004. Uptake of neutral metal complexes by a green alga: influence of pH and humic substances. Aust. J. Chem. 57, 931–936. Carrasco, L., Diez, S., Bayona, J.M., 2009. Simultaneous determination of methyl- and ethyl-mercury by solid-phase microextraction followed by gas chromatography atomic fluorescence detection. J. Chromatogr. A 1216, 8828–8834. Chen, J., Pehkonen, S.O., Lin, C.J., 2003. Degradation of monomethylmercury chloride by hydroxyl radicals in simulated natural waters. Water Res. 37, 2496–2504. Clarisse, O., Lotufo, G.R., Hintelmann, H., Best, E.P.H., 2012. Biomonitoring and assessment of monomethylmercury exposure in aqueous systems using the dgt technique. Sci. Total Environ. 416, 449–454. Clayden, M.G., Kidd, K.A., Wyn, B., Kirk, J.L., Muir, D.C.G., O’Driscoll, N.J., 2013. Mercury biomagnification through food webs is affected by physical and chemical characteristics of lakes. Environ. Sci. Technol. 47, 12047–12053. Cusnir, R., Steinmann, P., Bochud, F., Froidevaux, P., 2014. A DGT technique for plutonium bioavailability measurements. Environ. Sci. Technol. 48, 10829– 10834. Fernandez-Gomez, C., Bayona, J.M., Diez, S., 2012. Laboratory and field evaluation of diffusive gradient in thin films (DGT) for monitoring levels of dissolved mercury in natural river water. Int. J. Environ. Anal. Chem. 92, 1689–1698. Fernandez-Gomez, C., Bayona, J.M., Diez, S., 2014. Comparison of different types of diffusive gradients in thin films samplers for measurement of dissolved methylmercury in freshwaters. Talanta 129, 486–490. Fernandez-Gomez, C., Dimock, B., Hintelmann, H., Diez, S., 2011. Development of the DGT technique for Hg measurement in water: comparison of three different types of samplers in laboratory assays. Chemosphere 85, 1452–1457. Fernandez-Gomez, C., Drott, A., Bjorn, E., Diez, S., Bayona, J.M., Tesfalidet, S., Lindfors, A., Skyllberg, U., 2013. Towards universal wavelength-specific
photodegradation rate constants for methyl mercury in humic waters, exemplified by a Boreal lake-wetland gradient. Environ. Sci. Technol. 47, 6279–6287. Gardfeldt, K., Sommar, J., Stromberg, D., Feng, X.B., 2001. Oxidation of atomic mercury by hydroxyl radicals and photoinduced decomposition of methylmercury in the aqueous phase. Atmos. Environ. 35, 3039–3047. Gorski, P.R., Armstrong, D.E., Hurley, J.P., Krabbenhoft, D.P., 2008. Influence of natural dissolved organic carbon on the bioavailability of mercury to a freshwater alga. Environ. Pollut. 154, 116–123. Grandbois, M., Latch, D.E., McNeill, K., 2008. Microheterogeneous concentrations of singlet oxygen in natural organic matter isolate solutions. Environ. Sci. Technol. 42, 9184–9190. Hammerschmidt, C.R., Fitzgerald, W.F., 2006. Photodecomposition of methylmercury in an arctic Alaskan lake. Environ. Sci. Technol. 40, 1212– 1216. Hammerschmidt, C.R., Fitzgerald, W.F., 2010. 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Diffusive gradients in thin films for predicting methylmercury bioavailability in freshwaters after photodegradation C. Fernández-Gómez, J.M. Bayona, S. Díez ⇑ Environmental Chemistry Department, Institute of Environmental Assessment and Water Research, IDÆA-CSIC, E-08034 Barcelona, Spain
h i g h l i g h t s Assessment of DGT-labile MeHg during photodegradation was successfully tested. The action of sunlight seems not to alter the lability of MeHg. The kpd MeHg is mainly affected by DOM, which plays a key role in light attenuation. The quality of DOM, rather than the quantity, is key in the bioavailability of MeHg.
a r t i c l e
i n f o
Article history: Received 21 November 2014 Received in revised form 12 February 2015 Accepted 25 February 2015
Handling Editor: X. Cao Keywords: Methylmercury Photodegradation DGT Bioavailable Polyacrylamide
a b s t r a c t Determination of the dissolved-bioavailable fraction of methylmercury (MeHg) and its degradation pathways in freshwaters deserve attention, to further our understanding of the potential risk and toxicity of MeHg. Since the photodegradation of MeHg is the most important known abiotic process able to demethylate MeHg, this study investigated the role of sunlight on MeHg bioavailability in freshwater environments. Experiments to calculate photodegradation rate constants of MeHg in different types of freshwater in combination with experiments to distinguish the labile fraction of MeHg after being exposed to sunlight were performed. The ability of diffusive gradients in thin films based on polyacrylamide (P-DGT) to assess DGT-labile MeHg during photodegradation was successfully tested. First order photodegradation rate constants (kpd) of bioavailable MeHg determined in five different types of waters with different amount of dissolved organic matter (DOM), were in the range 0.073–0.254 h1, confirming previous findings that once there is DOM in solution, which would favour the photodegradation process, the kpd is mainly affected by light attenuation. Simulated sunlight seems not to alter the lability of MeHg, although photodegradation processes may decrease the concentrations of MeHg, contributing to reduce the amount of bioavailable MeHg (i.e. MeHg uptake by DGT). However, the quality of DOM, rather than the quantity, plays an important role in the bioavailability of MeHg in freshwater. Ó 2015 Elsevier Ltd. All rights reserved.
1. Introduction Methylmercury (MeHg) is one of the most widespread waterborne contaminants and tends to bioaccumulate and biomagnify throughout aquatic trophic chains (Clayden et al., 2013; Lavoie et al., 2013). Therefore, determination of MeHg degradation rates in aqueous matrices and identification of the factors involved in its bioavailability (i.e., available for uptake by biota) are crucial for a better understanding of the potential risk and toxicity of mercury. It has been shown that MeHg can be demethylated in the water column of lakes by the action of microorganisms (Matilainen and Verta, 1995; Schaefer et al., 2004), but recently, ⇑ Corresponding author. Tel.: +34 93 4006100; fax: +34 93 2045904. E-mail address: [email protected] (S. Díez). http://dx.doi.org/10.1016/j.chemosphere.2015.02.060 0045-6535/Ó 2015 Elsevier Ltd. All rights reserved.
abiotic processes driven by sunlight have received increasing attention (Sellers et al., 1996; Lehnherr and Louis, 2009; Hammerschmidt and Fitzgerald, 2010; Black et al., 2012). The photodegradation of MeHg is the most important known abiotic process capable of demethylating MeHg and has been suggested to be the principal sink for MeHg in lakes (Sellers et al., 2001; Lehnherr and Louis, 2009) consuming up to 80% of MeHg total input (Hammerschmidt et al., 2006; Hines and Brezonik, 2007). Visible and, more efficiently, ultraviolet (UV) radiation (Lehnherr and Louis, 2009; Black et al., 2012) are able to decompose MeHg, mainly by indirect photolysis, involving reactive oxygen species (ROS, e.g., OH, O2H, and 1O2) (Suda et al., 1993; Gardfeldt et al., 2001; Chen et al., 2003; Zhang and Hsu-Kim, 2010a) formed by sunlight irradiation of water and its dissolved organic matter (DOM; i.e. chromophoric groups in humic substances, such as
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quinones) (Alegria et al., 1997; Grandbois et al., 2008) or dissolved redox active inorganic components such as nitrate (Zepp et al., 1987) and iron (Hammerschmidt and Fitzgerald, 2010). Apart from its role in the formation and protection of ROS, DOM has been shown to attenuate UV light (Morris et al., 1995; Williamson et al., 1996) and thereby the MeHg photolysis process. Other variables such as solar radiation intensity and spectral composition, water constituents other than DOM (e.g., salinity, nitrate, or photoreactive trace metals), have been suggested to play a role in the photodegradation of MeHg (Zepp et al., 1987; Sellers et al., 1996; Chen et al., 2003; Hammerschmidt and Fitzgerald, 2006, 2010; Lehnherr and Louis, 2009; Black et al., 2012). In freshwater environments, both inorganic mercury and MeHg can be partitioned between suspended particulate matter (SPM), dissolved and colloidal water phases. The last is associated with the DOM occurring in freshwater ecosystems (Babiarz et al., 2001; Hill et al., 2009). This influences the speciation and potentially the bioavailability of mercury. Furthermore, the effect of solar radiation on the equilibrium among the different Hg species and its role in their bioavailability are still unknown. Determining the bioavailability of mercury is also essential in order to assess its risk and potential effects on exposed biota. For this purpose, the technique of diffusive gradients in thin films (DGT) can be used. This technique is widely used for passive sampling of trace metals, and allows the determination of timeweighted average concentrations of bioavailable contaminant species (Cusnir et al., 2014). The DGT technique has been successfully used for in situ determination of kinetically labile metal species in aquatic systems, such as MeHg (Clarisse et al., 2012) and is an effective tool for the assessment of bioavailability of Hg species (Amirbahman et al., 2013). The principle of the DGT technique is based on the diffusion of the dissolved species through a diffusive layer and their accumulation in an ion-exchange resin. A hydrogel and a membrane filter are commonly used as the diffusive layer and the resin is incorporated into a polyacrylamide gel. These three layers are enclosed and sealed in a small plastic device, so that only the membrane is exposed to the deployment solution. The timeaveraged concentration of the metal in the solution, C, can be calculated according to the Fick’s first law of diffusion as:
C¼
M Dg DAt
ð1Þ
where D is the diffusion coefficient of the metal in the diffusive layer, t is the deployment time, A the exposure surface area, and Dg the thickness of the diffusive layer. The mass (M) of the analyte accumulated by the resin is experimentally measured and provides the average labile metal concentration during the exposure time. The objective of this work was to study the photodegradation of MeHg in natural waters and to assess whether this process affect the bioavailability of MeHg. To accomplish this, we combined the calculated experimental photodegradation rate constants (kpd) in different types of freshwater with the bioavailable fraction of MeHg assessed by the DGT technique. Differences in photodegradation rates among the different types of freshwater were evaluated in terms of DOM concentration and conductivity, whereas differences in MeHg bioavailability were explained based on DOM concentration and quality, electrical conductivity, and the equilibrium binding constants of the MeHg species present. 2. Materials and methods 2.1. DGT manufacturing An in-house manufactured polyacrylamide-based DGT (P-DGT) device was used for assessment of labile or bioavailable MeHg. Details of diffusive gel, resin gel preparation, and DGT device
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assembly have been outlined previously (Fernandez-Gomez et al., 2011, 2014). Briefly, the binding layer, also known as the resin gel, consisted of a 3-mercaptopropyl-functionalised silica gel embedded in a polyacrylamide gel (0.4 mm thick). On top of the resin gel was a 0.4 mm thick disk of polyacrylamide gel used as the diffusive gel that was covered by a 0.1 mm thick, 0.45 lm pore size Nylon filter membrane. The reagents and materials employed for the preparation of the DGT gels and for MeHg determination are described in previous works (Fernandez-Gomez et al., 2011, 2012). Thiourea (ACS, P99.0%), purchased from Fluka (Steinheim, Germany), and 37% fuming hydrochloric acid (ACS, ISO), purchased from Merck (Germany), were used to prepare an acid thiourea solution for MeHg elution from resin gels. Methylmercury chloride (CH3HgCl, 99%) was obtained from Strem (Newburgport, MA, USA), and dimethylmercury chloride ((CH3)2HgCl) from Alfa Aesar (Karlsruhe, Germany). Stock standard solutions were prepared at 1000 mg L1 (as Hg) in acetone and stored at 20 °C. Working solutions were prepared weekly by diluting the stock solutions with acetone to a range of 0.02–500 lg L1 (as Hg). 2.2. Apparatus Gas chromatography (GC) coupled to atomic fluorescence spectrometry via a pyrolytic reactor (GC-Py-AFS) was used for the analysis of MeHg in water samples and DGT resin gel eluates. The GC analysis was accomplished with a non-commercial system formed by a Thermo Trace GC ultra (Milan, Italy) gas chromatograph interfaced to an AFS Tekran Model 2500 (Toronto, Canada) detector via a pyrolyser (Hg-800, Rektorik R&D Chromatography, Meyrin, Switzerland). Details of GC analysis have been described previously (Carrasco et al., 2009). 2.3. Freshwater origin Five different types of water were used: MilliQ water (NaCl 0.01 M) (MQW); MilliQ water (NaCl 0.01 M) with organic matter (10 mg DOC L1) from a Nordic reservoir NOM (IHSS, 1R108N) (MQO); water from the Ebro River at Flix reservoir, Catalonia, Spain (FLX); and water taken at two sites along a freshwater wetland-lake gradient in Boreal Sweden. FLX water was collected at the Flix reservoir (41°140 N, 0°320 E) on the course of the Ebro River (annual mean = 255–424 m3 s1; watershed area 85 820 km2). The Flix reservoir is relatively small (area = 320 ha; volume = 11 hm3), with a very short water residence time of 0.15 d. The two Swedish sites were a dystrophic lake, Ängessjön (ANG; 64°20 5600 N 20°500 1700 E), and a mixed Sphagnum/Carex peatland-dystrophic lake with approximately 30% open water, Kroksjön (KSN; 63°570 800 N 20°380 1300 E). The ANG lake has an average water depth of about 2–2.5 m and more than 50 cm deep organic sediments. About 30% of the shores are covered by peatland and 70% by shallow mineral soils. The KSN site is dominated by floating mats and islands of Sphagnum/Carex peat mixed with small areas with open water of a maximum water depth of about 1 m. 2.4. Calibration of DGT units for MeHg measurements In order to measure MeHg in the five different waters described above, two MeHg solutions (1 lg L1 MeHgCl, 0.01 M NaCl), one with organic matter from a Nordic reservoir (IHSS, 1R108N) (10 mg L1) and the other one without, were prepared in 5 L amber glass bottles. The pH was adjusted to 7 using 1M NaOH or HCl. Afterwards, the solution was left to equilibrate overnight while being stirred with a magnetic stirring bar. Ten DGT devices of PDGT were submerged in each solution, whose temperature was
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controlled and kept at 25.5 ± 0.5 °C. Duplicates of DGT units were retrieved after 2, 4, 8, 14, and 24 h. A duplicate that was not deployed was used as the DGT blank (0 h). At each sampling interval, 5 mL of each MeHg solution were collected and acidified to 0.4% with HCl to monitor the MeHg concentration remaining in solution, and 10 mL were taken only from the DOM-containing solution to measure dissolved organic carbon (DOC). Both the DGT units and the water samples were stored in the fridge until analysis within one week. 2.5. MeHg photodegradation experiment All experiments were conducted in acid-washed 500 mL widemouth TeflonÒ fluorinated ethylene propylene (FEP) bottles (Thermo Scientific-Nalgene, Rochester, NY, USA). Two bottles were filled with every water type; one replicate was used for the photodegradation process and the other as dark control. Every bottle was filled to the top, to which was added 250 ng Hg (CH3HgCl in acetone). The bottle was shaken vigorously, wrapped with aluminium foil, and left to equilibrate overnight at room temperature. In the case of the Swedish waters, since they had very high DOC levels, they were diluted to roughly 10 mg L1 of DOC. The following morning, the bottle was unwrapped and placed in the SuntestÒ for 7.5 h. Before light exposure (0 h) and at different time points during the experiment (0.5, 1.5, 3.5, and 7.5 h), the pH, conductivity, and DOC were determined. Moreover, at these time points, 2 mL of water were taken, preserved from light in an aluminium foil-wrapped glass vial, and acidified to 0.4% with HCl. Then 20 lL of ethylmercury chloride (EtHgCl; 230 ng L1) were added as internal standard and the water was stored at 4 °C until analysis. At the end of the photodegradation experiment, the bottle was double wrapped with aluminium foil and kept in the fridge until the bioavailability experiment (Section 2.6). The same procedure was used for dark controls, except that the bottles were double wrapped with aluminium foil also when placed inside the SuntestÒ equipment and aliquots were taken only three times (0, 3.5, and 7.5 h). The photodegradation rate constants (kpd) were calculated from first-order decay kinetics as the slope of the negative linear relationship between the natural logarithm of the quotient between the concentration of MeHg at the sampling occasion and the initial concentration, ln([MeHg]/[MeHgi]), and the time of exposure to the xenon lamp (h). 2.6. DGT-labile MeHg experiment The bottles that were exposed to photodegradation and those used as dark controls were used for this experiment. Two P-DGT units were introduced into FEP bottles exposed (n = 5) and unexposed (n = 5) to sunlight. All bottles were double wrapped with aluminium foil and placed in the shaking incubator Innova 40R (New Brunswick Scientific, Edison, NJ, USA) under conditions of 25 °C and 30 rpm. After 23 h, the DGT units were retrieved and stored in the fridge (4 °C). To monitor the MeHg concentration in the water throughout the experiment, 2 mL aliquots of water were taken from each bottle before and after the DGT deployment. 2.7. Calculation of the photodegradation yield index The photodegradation yield index (PDY) was estimated by decoupling the global spectral irradiation in specific values. This calculated PDY is not an experimental value but it should be useful for estimating the relative amount of MeHg photodegraded in the middle of FEP bottles (i.e. the longitudinal axis of the cylindrical bottle). The overall MeHg PDY (PDYoverall) is the addition of the specific PDY at different wavelengths (UVA, UVB, and visible light) (Eq. (2)).
PDYoverall ¼ PDYVis þ PDYUVA þ PDYUVB
ð2Þ
For each wavelength, the PDY is calculated as the product of three fractions: downwelling irradiance in the middle of the bottle (fd), contribution in the photodegradation process (fp), and contribution in the spectrum of the light source (i.e. the lamp) used for the experiment (fs), as follows:
PDYVis ¼ fdVis fpVis fsVis PDYUVA ¼ fdUVA fpUVA fsUVA
ð3Þ
PDY UVB ¼ fdUVB fpUVB fsUVB The fd at different wavelengths (Vis, UVA and UVB) were calculated using the wavelength specific light attenuation coefficients (Kd) described by Morris et al. (1995) and the downwelling irradiance equation described by Williamson et al. (1996) (Eq. (4)), where z is the radius of the bottle.
fdVis ¼ exp½ðK dVis Þz fdUVA ¼ exp½ðK dUVA Þz
ð4Þ
fdUVB ¼ exp½ðK dUVB Þz The different fp were obtained from the wavelength specific first order photodegradation rate constants (kpd): 0.0023 ± 0.0002, 0.10 ± 0.024, and 7.2 ± 1.3 m2 E1, for kpdPAR, kpdUVA, and kpdUVB, respectively, resulting in a relative ratio of 1:43:3100 for the three constants, reported in a previous study (Fernandez-Gomez et al., 2013) (Eq. (5)).
fpVis ¼
1 ; 3144
fpUVA ¼
43 ; 3144
fpUVB ¼
3100 3144
ð5Þ
Finally, the fs was obtained from the global spectral distribution provided by the instrument supplier of the xenon lamp used in the experiment (50%, 5.6%, and 0.4% of visible light, UVA, and UVB, respectively) (Eq. (6))
fsVis ¼ 0:50;
fsUVA ¼ 0:056;
fsUVB ¼ 0:004
ð6Þ
2.8. Analytical methods 2.8.1. Elution of MeHg from DGT resin gels After being retrieved from the DGT assembly, every resin gel was placed in a glass vial using hydrochloric acid (10%)-cleaned tweezers. Two millilitres of a freshly prepared thiourea acid solution (1.3 mM thiourea, 0.1 M HCl) were added to each vial, which was wrapped with aluminium foil to prevent photodegradation, and left at room temperature for the elution to take place. Following 24 h of exposure, the vials were stored in the fridge (4 °C) until analysis (within a week). 2.8.2. MeHg determination A 250 lL aliquot of resin extract was placed in a 6 mL glass vial with a polytetrafluoroethylene (PTFE)-coated silicone rubber septum containing 3 mL of citric-citrate buffer (pH = 4.5–5). Ten microlitres of EtHgCl (230 pg mL1 as Hg) in acetone were also added to the vial as instrumental standard. Once the vial was covered, it was placed in a water bath at 70 °C and stirred at 1200 rpm using an ETS-D4 fuzzy thermometer and a digital RCT hot plate, both purchased from Ika Labortechnik (Staufen, Germany). Then, 100 lL of 1% aqueous solution of NaBPh4 were added to the mixture through the septum as a derivatising agent. After a 5-min derivatisation step, extraction was conduted for 27 min by headspace solid-phase microextraction (HS-SPME) using a 100 lm polydimethylsiloxane (PDMS) fibre. The final detection was carried out using a GC-Py-AFS. Three resin gels were spiked with 3 ng of MeHg in order to calculate the recovery of the method, which was 65% (RSD = 4%, n = 3). This value was later used to correct
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the results. In the case of the water samples, the methodology was basically the same but with two differences: the EtHgCl was added as internal standard right after taking the samples and before storing them to correct for any possible loss of MeHg during storage or due to adsorption on the vial walls, apart from detector variations; and the aliquot of the sample for HS-SPME was 1 mL. Every time the instrument was switched on to run a set of samples, the repeatability of the methodology was tested by analysing the same sample, which yielded an RSD of less than 5% (n = 3), proving the robustness of the analytical procedure. A six-point calibration curve was performed in the range from 29 to 1180 pg of MeHg (as Hg). The limit of detection (LOD) was calculated as the mean of the method blanks plus three times their standard deviation, and the limit of quantification (LOQ) as the mean of the method blanks plus 10 times their standard deviation, resulting in 5.8 and 19 pg, respectively. 2.8.3. Statistical analysis The difference between the DGT-labile MeHg concentrations in the water that underwent photodegradation and in the dark control was analysed with a Mann–Whitney U test. Comparison of the diffusion coefficients of MeHg with and without DOM was performed by a two-sample t-test. The significance level was set at p 6 0.05. Statistical analysis was performed using the SPSS 15.0 for Windows software (SPSS Inc., Chicago, IL, USA). 3. Results and discussion 3.1. Calibration of DGT units for MeHg measurements The results of the calibration showed that DGT based on polyacrylamide accumulated MeHg in proportion to exposure time both in the presence and the absence of DOM (Fig. 1). The diffusion coefficient of MeHg in the diffusive layer can be calculated from the slope (s) of the relationship between the mass of MeHg accumulated by the DGT units and the deployment time. The diffusion coefficients of MeHg in the diffusive layer in the absence and in the presence of DOM in P-DGT were significantly different, with values of 2.54 106 and 1.67 106 cm2 s1, respectively, at 26 °C. These values agree with those that had been obtained previously (Fernandez-Gomez et al., 2014). Values of diffusivity for MeHg in the DGT diffusive layer in the absence and presence of DOM were not so different from each other as for inorganic Hg determined in a previous work (Fernandez-Gomez et al., 2011) which might suggest that MeHg-DOM complexes are more labile than Hg(II)-DOM complexes and have a higher dissociation rate
Fig. 1. Time-series experiment. Mass of MeHg ([MeHg]i = 1 lg L1, 0.01 M NaCl) accumulated in the resin at different deployment times, in the absence (MilliQ) and in the presence of DOM (10 mg L1).
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constant. Moreover, this result highlights the importance of calibrating the DGT units in a solution with similar characteristics (e.g., ion strength, DOM, etc.) to those where DGT would be deployed. 3.2. MeHg photodegradation experiment As shown in Fig. 2, for all five types of water, the pH remained fairly stable during the experiment. However, the DOC concentration decreased slightly due to DOM photolysis (Mopper et al., 1991; Hongve, 1994). In Fig. 3, it can be observed that for all the waters the MeHg concentration diminished when the bottles were exposed to the xenon lamp. On the contrary, the dark controls presented constant MeHg levels throughout the experiment. The five aqueous matrices used in this experiment covered a wide range of pH (5.2–8.3). However, as reported previously pH does not significantly affect the kpd of MeHg (Sellers et al., 1996; Chen et al., 2003). Concerning the electrical conductivity, previous studies (Zepp et al., 1987; Chen et al., 2003) observed an enhancement in kpd of MeHg with increasing chloride concentration, and proposed that it was due to the aqueous oxidation of chloride by hydroxyl radicals, producing chlorine radicals that may attack the Hg–C bond. Taking these findings into consideration, the higher kpd in FLX water (Table 1) might be explained by its high electrical conductivity (1010 lS cm1) in comparison with the Swedish waters (ANG and KSN, 11–40 lS cm1). However, MQO and MQW waters, despite also having high conductivity (1150 lS cm1), presented lower MeHg kpd rate constants (0.106 and 0.073 h1, respectively). The reason for the poorer linearity (R2 = 0.89) of the kpd rate in MQW in contrast to the rest of the water types (Fig. 3) is likely the lack of organic matter, which has been shown to contribute to MeHg photodecomposition by means of oxygen reactive intermediates (OH and 1O2) (Alegria et al., 1997; Chen et al., 2003; Zhang and Hsu-Kim, 2010b) generated by the interaction of sunlight with DOM. Unlike previous works,
Fig. 2. Changes in DOC and pH during the course of the MeHg photodegradation experiment.
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(a)
0.2 0 -0.2
PD
-0.4
Dark y = -0.0729x -0.0143 R² = 0.8871
-0.6 -0.8
0
2
4
6
ln ([MeHg]/[MeHgi])
ln ([MeHg]/[MeHgi])
188
8
(b)
0.2 0 -0.2
PD
-0.4
Dark
-0.6
y = -0.1057x + 0.0466 R² = 0.9715
-0.8 -1
0
2
(c)
0 -0.4 PD
-0.8
Dark
-1.2 y = -0.2538x -0.0812 R² = 0.9903
-1.6 -2
0
2
4
6
8
ln ([MeHg]/[MeHgi])
ln ([MeHg]/[MeHgi])
0.4
0.2
6
8
(d)
0 -0.2 PD
-0.4
Dark
-0.6 y = -0.1178x + 0.0447 R² = 0.9881
-0.8 -1
0
Time (h)
ln ([MeHg]/[MeHgi])
4
Time (h)
Time (h)
2
4
6
8
Time (h)
(e)
0.2 0 -0.2
PD
-0.4
Dark
-0.6
y = -0.1119x + 0.0689 R² = 0.9734
-0.8 -1
0
2
4
6
8
Time (h) Fig. 3. Relationships between ln([MeHg]/[MeHgi]) and light exposure time (h). The kpd full spectrum (m2 E1) was determined as the negative of the slope of the linear relationship. (a) MilliQ water (NaCl 0.01 M) (MQW), (b) MilliQ water (NaCl 0.01 M) with organic matter (MQO), (c) water from the Ebro River (FLX), (d) water from Ängessjön lake (ANG), and water from, (e) Kroksjön peatland-lake (KSN). PD: photodegradation experiment; Dark: dark controls.
Table 1 Water quality parameters, lability and bioavailability information. Initial DOC concentration, average pH along the experiment, conductivity, MeHg photodegradation rate constant (kpd), ratios between the MeHg mass accumulated by one DGT unit (M) and the average MeHg concentration (measured directly) in the deployment water during sampling (C), and percentages of bioavailable MeHg in the dark control and in the water exposed to artificial sunlight. Water type
[DOC] (mg L1)
pH
kpd (h1)
Cond.(lS cm1)
M/C (Dark) (cm3)
M/C (Light) (cm3)
% Bioav. (Dark)
% Bioav. (Light)
ANG KSN FLX MQO MQW
10.0 12.3 2.1 9.3 60.010
6.7 5.7 8.3 5.4 5.2
0.118 ± 0.009 0.112 ± 0.013 0.254 ± 0.014 0.106 ± 0.013 0.073 ± 0.018
40 11 1010 1150 1150
4.6 ± 0.4 4.4 ± 0.4 6.7 ± 1.0 8.3 ± 1.5 6.9 ± 1.8
4.8 ± 0.8 5.1 ± 0.2 6.1 ± 0.5 7.8 ± 2.1 7.8 ± 1.0
52 ± 5 50 ± 5 76 ± 11 95 ± 17 53 ± 14
55 ± 9 58 ± 3 70 ± 5 88 ± 24 60 ± 8
which suggested that photodegradation of MeHg hardly occurs in the absence of DOM (Chen et al., 2003; Zhang and Hsu-Kim, 2010a) it has been recently reported that MeHg can be degraded by sunlight when it is dissolved in plain ultra-pure water in the presence of chlorides (Sun et al., 2013). In the same manner, we observed slight photodecomposition of MeHg even in our MilliQ water with NaCl 0.01 M in solution (Fig. 3a, kpd-MQW = 0.073 h1). The lower kpd found for MQO and boreal waters (i.e., AGN and KSN) in comparison with FLX could be attributed to a higher DOM concentration, which attenuated more intensely the light penetrating the bottle. This might confirm previous findings (Fernandez-Gomez et al., 2013) that unless the MeHg degradation reaction is not limited by the generation of free radicals or by the complexation of MeHg by thiol groups, the main effect of DOM on MeHg photodegradation seems to be light attenuation. 3.3. Photodegradation yield index In order to better explain the difference in the kpd values between FLX and DOM-dominated waters, the PDY index was
calculated. Eqs. (3)–(6) were employed to calculate fp, fs, and fd values, and subsequently, PDY of MeHg for every light spectral range (UVB, UVA and visible light) as it is shown in Table 2. Then, it was possible to compare two waters by matching the results obtained in the Eq. (2), e.g. PDYoverall. For example, the obtained PDY values for MQO and FLX waters were 0.0021 and 0.0041, respectively (Table 2). Thus, the PDYMQO:PDYFLX ratio was 1:2.0. This result can provide the most likely explanation of the difference between the kpd of those two waters, since the kpd-MQO:kpdFLX ratio was 1:2.4, quite similar to the PDYMQO:PDYFLX ratio. Therefore, these results suggest that the main variable affecting MeHg photodegradation might be the DOC concentration, which plays an important role in attenuation of the different light wavelengths. Obviously, there were other variables that were not analysed that could affect the photodecomposition process. Analogous results were obtained among the rest of our waters (PDYANG:PDYFLX = 2.1 vs. kpd-ANG:kpd-FLX = 2.2; and PDYKSN:PDYFLX = 2.7 vs. kpd-KSN:kpd-FLX = 2.3). Concerning the statistical comparison (t-student, a = 0.05) of the obtained kpd values, no significant difference was found, except
C. Fernández-Gómez et al. / Chemosphere 131 (2015) 184–191 Table 2 Calculation of MeHg photodegradation yield index (MeHg PDY). Fraction of contribution in the photodegradation process (fp), fraction of contribution in the spectral composition of the lamp (fs), fraction of downwelling irradiance reaching the middle of the bottle with water (fd), and PDY of MeHg for every light spectral range (UVB, UVA and visible light). UVB
UVA
Vis
fp fs
0.986 0.004
0.014 0.056
0.0003 0.500
fd FLX (2.1 mg OC L1) MQO (9.3 mg OC L1) ANG (9.9 mg OC L1) KSN (12.3 mg OC L1)
0.825 0.362 0.336 0.249
0.924 0.643 0.622 0.543
0.982 0.921 0.917 0.897
FLX MQO ANG KSN
PDYUVB
PDYUVA
PDYVIS
MeHg PDYoverall
0.0033 0.0014 0.0013 0.0010
0.0007 0.0005 0.0005 0.0004
0.0002 0.0001 0.0001 0.0001
0.0041 0.0021 0.0019 0.0015
in the case of FLX versus the rest of the waters (p 6 0.0008), and in the case of ANG versus SKM (p = 0.02). 3.4. Lability of MeHg in the different types of water As mentioned before, to accurately determine the labile fraction of Hg in water by DGT, a diffusion coefficient (D, Eq. (1)) obtained under similar conditions to those in the water in which the DGT units were deployed should be used. However, in this case we studied five different types of water with different characteristics, which made it difficult to have a DGT calibration (i.e., D determination) for each of them. Thus, to rigorously compare the lability of MeHg in those waters, we worked with the ratio between the MeHg mass accumulated by one DGT unit (M) and the average MeHg concentration (measured directly) in the tested water during sampling (C) (Table 1). In this manner, we did not have to deal with problems derived from the use of the same D for all the different waters. The results were subjected to a Mann–Whitney test to statistically compare the MeHg lability indicator (M/C) in the water that underwent photodegradation and in the dark control. The test showed no significant differences between the two treatments in any of the five types of water. Therefore, we can conclude that the action of sunlight seems not to change the lability of MeHg. Aside from the influence of solar radiation on MeHg lability, the differences among the five waters can be discussed. The high values of the equilibrium binding constants (log K) for the association of MeHg with thiol-containing compounds or humic substances (log KMeHg-DOM = 12.6–16.9) (Reid and Rabenstein, 1981; Hintelmann et al., 1997; Amirbahman et al., 2002; Qian et al., 2002) in comparison with the equilibrium binding constant of MeHgCl (log KMeHgCl = 5.25), at similar concentrations to those studied here, suggest that MeHg associates primarily with the thiol groups in the DOM. Thus, in freshwater systems, when DOM concentration is sufficiently high, the MeHg-DOM complex should be more abundant than the MeHgCl complex. Black et al. estimated that DOM bound >97% of the MeHg at a salinity of 25‰ and DOM concentration of 1.5 mg OC L1 when MeHg concentration was 0.25–1.25 ng L1. Therefore, under our experimental conditions MeHg-DOM binding should also be predominant. However, changes in the complexation of MeHg by DOM and chloride should be a function of DOM concentration, salinity, and pH. Thus, the lowest M/C ratios found in ANG (4.6–4.8) and KSN (4.4–5.1) waters (Table 1) could be explained by their high DOM concentration
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(P10 mg DOC L1) and their low conductivity (40 and 11 lS cm1, respectively), and hence by the prevalence of MeHgDOM complexes in solution. These complexes are bigger in size than MeHg+, which makes them diffuse at a slow rate through the DGT diffusive layer, and they dissociate with more difficulty within the diffusive gel. Consequently, since a smaller amount of MeHg is bound to the thiol groups in the resin gel, a minor proportion of MeHg will be measured. A similar phenomenon happens also when working with living organisms. Some authors stated that an increase in DOM concentration led to a decrease in MeHg bioavailability, as tested in algae, seston and hydropsychids, or bacteria (Gorski et al., 2008; Tsui and Finlay, 2011; Luengen et al., 2012; Ndu et al., 2012) as the MeHg-DOM complexes are too large or hydrophilic to pass through the cell membrane. Furthermore, the slightly higher M/C ratio in FLX water (6.1–6.7) could be accounted for by the same behaviour. This water had a lower DOM concentration (2 mg DOC L1) and a notably higher conductivity (1010 lS cm1) than the others. Therefore, there could be a higher amount of MeHgCl, which is more easily dissociable than MeHg-DOM complexes. Concerning MQO and MQW waters, the difference between them is that MQO had a DOM concentration of 9.3 mg L1 and MQW was DOM-free. Although log K was substantially higher for MeHg associated with humic substances than for MeHgCl, MQO presented a similar MeHg lability (7.8–8.3) as MQW (6.9–7.8). Again, the presence of organic matter seemed not to affect MeHg lability significantly, as shown by the similar values of diffusivity for MeHg in P-DGT in the absence and presence of DOM. On the other hand, in order to estimate the fraction of MeHg that could be available to biota in all the water types, the following approach was taken. The percentage of ‘bioavailable MeHg’ (Table 1) was calculated as the quotient between the DGT-labile MeHg (measured with DGT) and the average of the MeHg concentration (measured directly in the water before and after DGT deployment). The DGT-labile MeHg concentration was calculated using D of MeHg in the presence of DOM for all the waters except for MQW, for which D in the absence of DOM was employed. As stated before, the D used for the DGT calculations should be obtained from a calibration performed in water with similar or identical characteristics to those of the water in which the DGT units were deployed, if an accurate labile (and then, proxy for bioavailable) MeHg concentration determination is desired. The high percentage of bioavailable MeHg in MQO (88–95%) versus that in MQW (53–60%) does not match our previous claim, nevertheless it might be explained by Pickhardt and Fisher’s (Pickhardt and Fisher, 2007) findings, that showed that DOC may enhance uptake. These authors observed that MeHg bioaccumulation in eukaryotic cells was greater in high DOC water. This uptake enhancement could occur through association of DOM with cell surfaces, which may facilitate uptake of Hg species into cells and/or promote transport by influencing the permeability of the cell membrane (Vigneault et al., 2000; Boullemant et al., 2004). A similar phenomenon might occur in the DGT membrane filter and diffusion gel.
4. Conclusions Our findings show that the DGT technique was successful to monitor MeHg photodegradation. Values of diffusivity for MeHg in the P-DGT in the absence and presence of DOM suggest that MeHg-DOM complexes are more labile than Hg(II)-DOM complexes. Concerning the photodegradation study, it can be concluded that as long as there is organic matter in the solution, which favours the photodegradation process, the kpd of MeHg is mainly affected by the light attenuation effect by DOM. A new
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and simple index has been developed and successfully employed to calculate a rough and relative amount of MeHg photodegraded in the middle of the bottle, the so-called, photodegradation yield index, PDY. This index is very useful for comparing two waters by matching this index with the ratio of its kpd and for confirming that DOM is key in the photodegradation rate for MeHg in DOM (humic)-dominated waters. On the other hand, the action of sunlight seems not to change the lability of MeHg and, therefore, its bioavailability; however, the DOM quality could be the main variable affecting it. Moreover, photodegradation resulted in decreased concentrations of MeHg, and thus, should decrease the amount of MeHg that is bioavailable, or in more precise wording, taken up by the DGT sampler. Future works should be focused on studying the influence of the type of DOM or humic substances present in the water on the bioavailability of MeHg, since several studies have also suggested that DOM composition may determine site-specific bioavailability of MeHg. Acknowledgments Financial support was provided by the Spanish Ministry of Economy and Competitiveness, MINECO (Project CTQ201125614). CFG would like to express her gratitude to Spanish National Council of Research (CSIC) and European Social Fund (ESF) for a JAE-predoc fellowship. Special thanks are also due to Ulf Skyllberg’s group, for supplying some water samples. We acknowledge the comments of anonymous referees, which helped us to improve the manuscript. References Alegria, A.E., Ferrer, A., Sepulveda, E., 1997. Photochemistry of water-soluble quinones. Production of a water-derived spin adduct. Photochem. Photobiol. 66, 436–442. Amirbahman, A., Massey, D.I., Lotufo, G., Steenhaut, N., Brown, L.E., Biedenbach, J.M., Magar, V.S., 2013. Assessment of mercury bioavailability to benthic macroinvertebrates using diffusive gradients in thin films (DGT). Environ. Sci. Proc. Impacts 15, 2104–2114. Amirbahman, A., Reid, A.L., Haines, T.A., Kahl, J.S., Arnold, C., 2002. Association of methylmercury with dissolved humic acids. Environ. Sci. Technol. 36, 690–695. Babiarz, C.L., Hurley, J.P., Hoffmann, S.R., Andren, A.W., Shafer, M.M., Armstrong, D.E., 2001. Partitioning of total mercury and methylmercury to the colloidal phase in freshwaters. Environ. Sci. Technol. 35, 4773–4782. Black, F.J., Poulin, B.A., Flegal, A.R., 2012. Factors controlling the abiotic photodegradation of monomethylmercury in surface waters. Geochim. Cosmochim. Acta 84, 492–507. Boullemant, A., Vigneault, B., Fortin, C., Campbell, P.G.C., 2004. Uptake of neutral metal complexes by a green alga: influence of pH and humic substances. Aust. J. Chem. 57, 931–936. Carrasco, L., Diez, S., Bayona, J.M., 2009. Simultaneous determination of methyl- and ethyl-mercury by solid-phase microextraction followed by gas chromatography atomic fluorescence detection. J. Chromatogr. A 1216, 8828–8834. Chen, J., Pehkonen, S.O., Lin, C.J., 2003. Degradation of monomethylmercury chloride by hydroxyl radicals in simulated natural waters. Water Res. 37, 2496–2504. Clarisse, O., Lotufo, G.R., Hintelmann, H., Best, E.P.H., 2012. Biomonitoring and assessment of monomethylmercury exposure in aqueous systems using the dgt technique. Sci. Total Environ. 416, 449–454. Clayden, M.G., Kidd, K.A., Wyn, B., Kirk, J.L., Muir, D.C.G., O’Driscoll, N.J., 2013. Mercury biomagnification through food webs is affected by physical and chemical characteristics of lakes. Environ. Sci. Technol. 47, 12047–12053. Cusnir, R., Steinmann, P., Bochud, F., Froidevaux, P., 2014. A DGT technique for plutonium bioavailability measurements. Environ. Sci. Technol. 48, 10829– 10834. Fernandez-Gomez, C., Bayona, J.M., Diez, S., 2012. Laboratory and field evaluation of diffusive gradient in thin films (DGT) for monitoring levels of dissolved mercury in natural river water. Int. J. Environ. Anal. Chem. 92, 1689–1698. Fernandez-Gomez, C., Bayona, J.M., Diez, S., 2014. Comparison of different types of diffusive gradients in thin films samplers for measurement of dissolved methylmercury in freshwaters. Talanta 129, 486–490. Fernandez-Gomez, C., Dimock, B., Hintelmann, H., Diez, S., 2011. Development of the DGT technique for Hg measurement in water: comparison of three different types of samplers in laboratory assays. Chemosphere 85, 1452–1457. Fernandez-Gomez, C., Drott, A., Bjorn, E., Diez, S., Bayona, J.M., Tesfalidet, S., Lindfors, A., Skyllberg, U., 2013. Towards universal wavelength-specific
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