Mercury removal using ground and calcined mussel shell

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JOURNAL OF ENVIRONMENTAL SCIENCES ISSN 1001-0742 CN 11-2629/X

Journal of Environmental Sciences 2013, 25(12) 2476–2486

www.jesc.ac.cn

Mercury removal using ground and calcined mussel shell Susana Pe˜na-Rodr´ıguez1 , Alipio Berm´udez-Couso1 , Juan Carlos N´ovoa-Mu˜noz1 , ´ Manuel Arias-Est´evez1 , Mar´ıa J. Fern´andez-Sanjurjo2 , Esperanza Alvarez-Rodr´ ıguez2 , Avelino N´un˜ ez-Delgado2,∗ ´ 1. Area de Edafolox´ıa e Qu´ımica Agr´ıcola. Departamento de Biolox´ıa Vexetal e Ciencia do Solo. Univ. Vigo. Facultade de Ciencias, Ourense, 32004 Ourense, Spain 2. Departamento de Edafolox´ıa e Qu´ımica Agr´ıcola. Univ. Santiago de Compostela. Escola Polit´ecnica Superior, campus univ. s/n, 27002 Lugo, Spain Received 08 February 2013; revised 28 March 2013; accepted 09 April 2013

Abstract We determined mercury retention on calcined and ground mussel shell, in presence and absence of phosphate, using batch and stirred flow chamber experiments. In batch experiments the calcined shell exhibited higher Hg adsorption, with good fitting to Freundlich equation (R2 : 0.925–0.978); the presence of phosphate increased Hg adsorption; mercury desorption was 13% or lower, diminishing up to 2% under the presence of phosphates. In stirred flow chamber experiments calcined shell retained more Hg than ground shells (6300 vs. 4000–5200 µmol/kg); Hg retention increased an additional 40% on calcined shell and up to an additional 70% on ground shells when phosphates were present; mercury desorption was quite similar in all shell types (20%–34%), increasing up to 49%–60% in ground shells when phosphates were present. The higher Hg adsorption on calcined shell would be related to its calcite and dolomite concentrations; mercury-phosphate interactions would cause the increase in Hg retention when phosphates are present. Data on Hg desorption suggest that Hg retention was not easily reversible after batch experiments, increasing in the stirred flow chamber due to convective flow. Calcined and ground mussel shells could be recycled removing Hg from water, with the presence of phosphates in solution improving efficacy. Key words: adsorption; desorption; heavy metals pollution; mercury; mollusc shell DOI: 10.1016/S1001-0742(12)60320-9

Introduction Mercury pollution is a matter of great concern, this fact causing Hg removal to be an environmental challenge with worldwide repercussion. For example, Hg has been detected in large quantities in forest soils (N´ovoa-Mu˜noz et al., 2008), and various studies have found Hg being dragged down-slope in areas where deforestation, forestry and/or other activities that disturb forest soils took place (Almeida et al., 2005; Munthe and Hultberg, 2004; Porvari et al., 2003; Schwesig et al., 1999). In such cases, it would be interesting having means to treat Hg polluted runoff. China is the most important world producer of the mussel Mytilus galloprovincialis Lamarck, with Galicia (NW Spain) placed second or third (Caballero et al., 2009; Pe˜na-Rodr´ıguez et al., 2010). In Galicia most of this mussel production is processed by the cannery industry, which generates an amount of waste mussel shell as great * Corresponding author. E-mail: [email protected]

as 65,000–94,000 tonnes per year, depending on mussel production (Barros et al., 2009; Pe˜na-Rodr´ıguez et al., 2010). This waste needs to be properly recycled in order to minimize risks of environmental pollution. Mussel shell has traditionally been partially recycled on agricultural soils, mainly due to its pH correction potential in acid environments, and it has shown to be of agronomic ´ value (Alvarez et al., 2012a, 2012b). Farmers in coastal areas have usually held the direct recycling of untreated shell to increase the pH of acid soils, but in the last decades some factories have implemented industrial processes leading to the treatment of waste shells to transform it into products rich in carbonates and oxides, achieving appropriate particle size and eliminating residual organic matter, undesirable odours and microorganisms. These factories wash, crush and sieve the shells, and sometimes also perform calcination. Recently the use of mussel shell has been postulated in water treatment, mainly to retain phosphate and certain

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Mercury removal using ground and calcined mussel shell

heavy metals, thus performing another way to effectively recycling this waste material. Regarding phosphate removal, greater efficacy has been shown when the calcite/aragonite ratio increases. The shell of some species is mainly constituted by aragonite, and only after calcination it shows higher percentage of calcite (Abeynaike et al., 2011; Currie et al., 2007). Other species may have shells that originally contain higher percentage of calcite, or even high levels of both CaCO3 polymorphs (Cubillas et al., 2005a, 2005b). Currie et al. (2007) showed that the use of calcined mussel shells eliminated around 90% of the phosphates in the water samples analysed, whereas the efficacy was around 40% when not-calcined mussel shells were used. Further, the ability of bivalve shells to accumulate heavy metals has been reported in several studies (Carrasco et al., 2008; Yap et al., 2011). In this sense, Du et al. (2011) demonstrated that the presence of aragonite and calcite phases in mollusc shells is effectively able to remove some heavy metals (Pb, Cd and Zn) from aqueous solutions. Contrary to that occurring with phosphates, it has been shown that the degree of retention of divalent cations in the shell is higher when it contains more aragonite than calcite (Cubillas et al., 2005a; K¨ohler et al., 2007; Prieto et al., 2003; Sakulkhaemaruethai et al., 2010). Prieto et al. (2003) showed high Cd2+ retention in aragonite of cockle (Cardiide family) shells, while the removal was low using calcite. Cubillas et al. (2005a) also found higher retention of Cd2+ in aragonite using mussel shell and shell from other bivalves, and K¨ohler et al. (2007) showed high efficacy of aragonite in the removal of dissolved Cd, Pb and Zn. Pe˜na-Rodr´ıguez et al. (2010) have published the only existing study reporting mussel shell efficacy to retain Hg from water, showing that high levels of Hg could be removed using calcined mussel shell, although this material contained 7% of aragonite even after calcination. These authors found higher Hg adsorption in one of the three shell batches assayed, namely that with the highest Fe and Al concentrations, this fact indicating that other characteristics than the aragonite/calcite ratio would be taken into consideration when studying the shell removal mechanisms for certain pollutants, as is the case of Hg(II). A waste material derived from mussel shell calcination (shell ash) has also shown high Hg retention potential, although being much more difficult to recycle than mussel shell due to its physical and chemical characteristics (SecoReigosa et al., 2012). Even with all this background, there were no previous publication determining the Hg removal potential of ground not-calcined mussel shell (which is produced at lower cost, expending much lesser energy than with calcined shell), nor studying the effects of the presence of phosphate in solution (presence clearly feasible in polluted waters) on the Hg removal potential of calcined and ground not-calcined mussel shell.

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The objective of this study is to evaluate the Hg removal potential of three kind of mussel shells (not-calcined finely ground, not-calcined coarsely ground, and calcined shell), either in the presence or in absence of phosphate. For this purpose, we used both batch and stirred flow chamber experiments to study these three kind of mussel shell samples, all of them having different concentrations of aragonite, calcite, Fe, Al and other elements that could affect retention and release of Hg.

1 Materials and methods 1.1 Materials Three batches of shell from mussel (Mytilus galloprovincialis Lamarck) were used in this study: two different batches of ground (not-calcined) mussel shell, fine (size < 1 mm) and coarse (size range 1–3 mm), from Abonomar S.L. (Illa de Arousa, Pontevedra province, Spain), and another batch of calcined shell, supplied by the factory Calizamar S.L. (Boiro, A Coru˜na province, Spain), where calcination takes place in a rotary kiln at 550°C for 15 min. The calcined mussel shell used here corresponds to the batch 3 of calcined mussel shell previously studied by Pe˜na-Rodr´ıguez et al. (2010), which showed high Hg(II) adsorption potential in batch type experiments. 1.2 Characterization of the materials All shell batches were cleaned in an ultrasonic batch for 30 min and then rinsed with double-distilled water. Thus, shell batches were oven-dry at 40°C for 24 hr and then grounded using an agate mortar (model “RM 100”, Retsch Company, Germany) until all sample passed through a 1mm mesh (fine mussel shell) or 3-mm mesh (coarse mussel shell). The pH values were measured with pH-meter (Crison micro-pH 2001, Spain) using an open cell electrode in 1/2.5 shell/distilled water suspensions after 10 min agitation, time that was required to attain the equilibrium. All mussel shells used had alkaline pH: 9.4 (fine ground shell), 9.1 (coarse ground shell), and 9.2 (calcined shell). The crystalline composition of the shells was elucidated by X-ray diffraction analysis (Philips PW1710 diffractometer, The Netherlands). The specific surface area was measured using a surface area and porosity analyser (Micromeritics model ASAP 2020, USA). Carbon and nitrogen were determined by elemental analysis using a Leco 2000 NC auto-analyzer (USA). The total concentrations of metals were determined by atomic absorption/emission spectroscopy, after digestion of the samples with nitric acid in a microwave oven (Milestone ETHOS 900 Microwave Labstation, USA) (Pe˜na-Rodr´ıguez et al., 2010). Specifically, 0.2 g of finely ground sample were treated with 6 mL of concentrated HCl (37%, density 1.19 kg/ L), 2 mL of concentrated HNO3 (65%, density 1.40 kg/L) and 1 mL of ultrapure water;

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Journal of Environmental Sciences 2013, 25(12) 2476–2486 / Susana Pe˜na-Rodr´ıguez et al.

once the digestion process was completed, the metals (Na, K, Ca, Mg, Al, Fe, Mn, Zn, Cu) were quantified by atomic absorption/emission spectroscopy, in a Perkin Elmer Optima 4300 DV (USA) spectrophotometer. For quality assurance and control (QA/QC) purposes, we used a certified reference material endorsed by the Canadian Certified Reference Materials Project (soil from calcareous till parent material, SO-3). Triplicate samples of the reference standard material were digested following the above-described method and the digests were analysed for total metal content. The metals recovery were in reasonable agreements with the certified values departing less than 5% for Na, K, Ca, Mg, Al and Fe and less than 10% for Mn, Zn and Cu. In addition, X-ray fluorescence (USC in-house dispersion spectrophotometer, Spain) was used to quantify P, Cl and S. Detection limits were 0.01 mg/L. A mercury analyser (MA-2000 Nippon Instruments, Japan) with a gold coated trap was used to determine the concentration of mercury in the samples, after thermal decomposition in a ceramic combustion tube; the mercury was detected in a double channel non-dispersive atomic fluorescence spectrometer by cold vapour atomic absorption (at a wavelength of 253.7 nm). The analyzer was calibrated using 0.01 mg/L and 0.1 mg/L of Hg stock solutions made from 1000 mg/L HgCl2 standard in 0.001% L-cysteine (Nacalai Tesque, Inc., Kyoto, Japan). The detection limit was 0.002 ng and the working range 0.002–1000 ng, both expressed as absolute amount of Hg. All determinations were made by triplicate and the analysis were repeated when the coefficient of variation exceeded 5%. For quality assurance and control (QA/QC), standard reference materials NCR-MESS-3 (estuarine sediment) and NCS DC73323 (soil) were measured at the beginning of each analytical run and repeated every ten samples. When the deviation of Hg values of standard materials was higher than 5%, the Hg-analyzer was recalibrated and the samples re-analyzed when satisfactory values of standards were achieved. All concentrations were expressed on an oven dry basis (105°C). All determinations were made by triplicate, and coefficients of variation were lower than 5%. 1.3 Hg adsorption and desorption 1.3.1 Batch test For the adsorption studies, 10 mL of a 0.01 mol/L NaNO3 solution containing a known concentration of Hg(II) as Hg(NO3 )2 (1.25, 2, 5, 6, 7, 10, 13, 18, 21 and 25 µmol/L Hg) were added to 200 mg of mussel shell in 15 mL borosilicate glass tubes. The pH of the Hg solutions used in batch test ranged from 3.45 to 4.85. The samples were shaken for 24 hr in an end-over-end (48 revolutions per minute (r/min)), then centrifuged at 4000 r/min (6167 ×g) for 10 min. As shown in a previous study (Pe˜na-Rodr´ıguez et al., 2010), a shaken period of 24 hr was enough to

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achieve equilibrium. The amount of Hg(II) adsorbed was calculated as the difference between the amount added and that recovered after 24 hr contact. After the shaking period, the resulting suspensions were centrifuged and the supernatant filtered through acid washed filter paper of 0.45 µm porous size discarding the first mL. Afterwards, Hg(II) in the supernatant was measured by formation of cold vapour and atomic absorption spectrophotometry. All of the experiments were carried out in triplicate. Freundlich (Eq. (1)) and Langmuir (Eq. (2)) isotherms were used to describe the adsorption behavior of Hg(II). These equations are expressed as follows: 1

X = KF C n X=

K L Xm C 1 + KL C

(1) (2)

where, X (µmol/kg) is the amount of solute retained per unit weight of adsorbent; C (µmol/L) is the equilibrium concentration of the solute remained in the solution; KF (Ln kg−1 µmol1−n ) and 1/n (dimensionless) are the Freundlich coefficients in equation 1; KL (L/µmol) is a constant related to the energy of adsorption; Xm (µmol/kg) is the maximum adsorption capacity of the sample. At the end of the adsorption period (24 hr), the centrifuged residues were weighed to determine the amount of occluded solution and re-suspended in 10 mL of a Hg(II)free 0.01 mol/L NaNO3 solution; the suspensions were shaken for 24 hr. Then, the supernatants were removed by centrifugation at 4000 r/min (6167 ×g) for 10 min, filtered (0.45 µm porous size filters) and their concentration of Hg and pH measured. The process was carried out for the different initial concentrations of Hg(II) (1.25, 12.5 and 25 µmol/L solutions). Mercury desorption data are expressed as percentages of previously adsorbed Hg(II). For this calculation, the amount of occluded Hg (estimated as the difference between the final and initial weights) was taken into account. 1.3.2 Influence of phosphate on Hg adsorptiondesorption Shell samples (200 mg) were added with 10 mL of 0.01 mol/L NaNO3 solutions containing 50 µmol/L Hg, and P concentrations (as NaH2 PO4 ) in the range 0.16 mmol/L to 16 mmol/L and, then, were shaken for 24 hr and centrifuged at 4000 r/min (6167 ×g) for 10 min. The pH of the Hg + P solutions used in this batch test ranged from 3.86 to 4.09. As previously, the supernatant was used to measure Hg concentration and pH; further, also P concentration was determined by means of spectrometric analysis at 660 nm (Bran+ Luebbe Autoanalyzer 3, Germany). When the adsorption process had finished, desorbed Hg was measured in those samples where 0.16, 8 and 16 mmol/L P had been added. The analytical procedure was similar to that of the samples without P.

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Mercury removal using ground and calcined mussel shell

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1.3.3 Stirred-flow chamber experiments Experiments to determine Hg(II) retention and release were carried out in a stirred-flow chamber whose volume was 1.5 cm3 . Stirring was provided by a PTFE-coated magnetic bar (3 mm × 1 mm) that was spun at 400 r/min to provide a constant flow. The experimental procedure consisted of placing 200 mg of shell sample in the device and passing a 0.01 mol/L NaNO3 solution containing a known concentration of Hg(II) (25 µmol/L) through the chamber. The flow rate was 0.6 mL/min, and 3 mL volumes were collected in glass tubes (maintaining the flow for approximately 5 min per tube). Experimental data obtained were fitted to a first-order mathematical model considering just one adsorption site (Eq. (3)).

dqd = kd (q0 − qd ) dt

(4)

where, kd (min−1 ) is its desorption kinetic constant; q0 (µmol/kg) is the maximum quantity of Hg that could be desorbed under the experimental conditions; qd (µmol/kg) is the quantity of Hg desorbed from the shell samples. 1.4 Fitting to adsorption models The fitting to the various adsorption models was performed by means of SPSS 18.0.

2 Results 2.1 Characteristics of the samples

dqs = ks (qmax − qs ) dt

Table 1 shows the general chemical composition of the calcined mussel shell (sample a), and finely ground mussel shells (sample b) and coarsely ground mussel shell (sample c) used in this study. Sample b exhibits higher Al, Fe, Mn, Zn, P and S concentrations, followed by sample a, which is the sample with the highest concentrations of K, Ca and Mg. Sample c shows the highest N, Cl and Na concentrations. The patterns of X-ray diffraction analysis showed that samples b and c have aragonite and calcite concentrations higher than 40%; sample c contains no quartz and shows traces of halite. The sample a has lower percentage of aragonite than samples b and c, but its calcite concentration is above 61% and it is also characteristic a notable concentration of dolomite (29%). The surface area of the shells was similar for all three shell samples, although slightly higher in sample b (1.40 ±

(3)

where, dqs /dt (µmol/kg min) is the Hg adsorption rate; ks (min−1 ) is a kinetic constant; qmax (µmol/kg) is the maximum capacity for Hg adsorption under the experimental conditions; qs (µmol/kg) is the quantity of Hg retained into the shell. The total adsorption period was 400 min, after which desorption was carried out with a Hg(II)-free 0.01mol/L NaNO3 solution at the same flow rate and time. The Hg(II) concentrations in the samples were measured by atomic absorption spectrometry by formation of cold vapour. The adsorption-desorption process was further repeated adding 3 mmol/L P. Then, Hg, P and pH were measured. Data obtained for Hg desorption kinetics were fitted to a first-order kinetic model (Eq. (4)).

Table 1 Chemical composition of different materials Shell sample

Ctotal (g/kg)

S (g/kg)

N (g/kg)

Cl (g/kg)

Na (mg/kg)

K (mg/kg)

Ca (mg/kg)

Calcined mussel shell (a) Finely ground mussel shell (b) Coarsely ground mussel shell (c)

128 114 127

2.1 3.4 2.2

2.6 2.1 3.6

7.3 5.4 11.0

5326 5174 5508

503 202 81

399439 280168 298086

Shell sample

Mg (mg/kg)

Al (mg/kg)

Fe (mg/kg)

Mn (mg/kg)

Zn (mg/kg)

Cu(mg/kg)

P (mg/kg)

Calcined mussel shell (a) Finely ground mussel shell (b) Coarsely ground mussel shell (c)

1968 981 1021

374 433 94

602 1855 245

14 49 15

17 28 8

4 8 7

890 1018 656

a

b

×1,000 Fig. 1

10 μm

c

10 μm

10 μm

SEM micrographs of the three kind of shell samples. (a) calcined shell; (b) fine ground shell; (c) coarse ground shell.

Journal of Environmental Sciences 2013, 25(12) 2476–2486 / Susana Pe˜na-Rodr´ıguez et al.

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0.05 m2 /g) than in sample c (0.99 ± 0.04 m2 /g), and sample a (1.07 ± 0.05 m2 /g). Figure 1 shows scanning electron microscopy (SEM) micrographs of the three kind of mussel shells, evidencing morphological differences among samples. In this sense, calcined shell shows platy lamellar formations (clearly more relevant than that found in ground shell samples), similar to those previously found by Pe˜na-Rodr´ıguez et al. (2010) in calcined mussel shell from the same factory. 2.2 Batch experiments

1200 1000

Sample a

600 400 200

Adsorbed Hg (µmol/kg)

0.4 0.6 0.8 Equilibrium Hg (µmol/L)

1.0

0.4 0.6 0.8 Equilibrium Hg (µmol/L)

1.0

Sample b

800 600 400 200 0 0.0

Adsorbed Hg (µmol/kg)

0.2

1200 1000

0.2

1200 1000

Sample c

800 600 400 200 0 0.0

Freundlich parameters 1/n (dimensionless) KF (Ln /(kg·µmoln−1 ))

R2

a b c

1.11 ± 0.18 6.30 ± 0.70 1.06 ± 0.078

0.925 0.962 0.978

5205 ± 1581 1178 ± 72 890 ± 25

the calculation of the constants. The KF parameter follows the sequence: sample a > sample b > sample c. The n parameter was not different to 1 for samples a and c, while it was much higher for sample b, in accordance with the shape of the adsorption curve (Fig. 2). KF values indicate that sample a is the one with the highest Hg adsorption (KF = 5205), followed by sample b (finely ground shell) (KF = 1178), and sample c (coarsely ground shell) (KF = 890). Figure 3 shows that Hg adsorption increases after a given amount of P is added and then, no further increase or decrease is observed in Hg retention in spite of the raise in P addition. This behaviour is similar for all three shell samples and thus, Hg adsorption in the presence of the lowest P addition increased 54% for sample a, 49% for sample b, and 38% for sample c, as compared with those values observed in absence of P. Concretely, the addition of P concentrations 6 40 µmol/L increases Hg adsorption (which any case is always higher than 95%), while higher P concentrations does not affect it significantly. 2.2.2 Hg desorption

800

0 0.0

Hg adsorption parameters for samples a, b, c

Sample

Batch type experiments showed low and similar desorption percentages for all three shells (Table 3), although desorption was slightly lower for sample a (calcined) than for samples b and c (ground shells), especially when adding high Hg concentrations (in fact, the amount of Hg desorbed increases as a function of the increase in the initial Hg concentration). The concentrations of Hg desorbed were in the range 7 to 37 µmol/kg for sample a, 4 to 85 µmol/kg for sample b, and 7 to 71 µmol/kg for sample c. These results, when expressed in percentage referred to the previously adsorbed Hg, are very similar for all three shell samples (Table 3). The addition of phosphate drastically diminishes Hg desorption in all three shell samples. The Hg desorption 3000 Adsorbed Hg (µmol/kg)

Adsorbed Hg (µmol/kg)

2.2.1 Hg adsorption Figure 2 shows the adsorption curves for the three kind of mussel shell samples. Sample a shows high adsorption affinity, with a trend to linearity. Sample b has S shape curve, with slight adsorption at the initial low Hg concentrations, and later changes in slope as the Hg load increases. Sample c shows an adsorption curve with a trend to linearity as sample a, but with lower slope. Experimental data were fitted to the Langmuir and Freundlich models, showing good fitting to the Freundlich equation, with R2 values in the range 0.925 to 0.978 (Table 2), but with poor fitting to the Langmuir equation due to the existence of too high error values associated to

Table 2

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2500 2000 1500 1000 500 0 0

0.2

0.4 0.6 0.8 1.0 Equilibrium Hg (µmol/L) Fig. 2 Hg adsorption curves for the shell samples a, b, and c. Symbols represent mean values of triplicates with all standard error below 15%.

Sample a Sample b Sample c

500 1000 1500 2000 2500 3000 3500 4000 4500 5000 Added P (µmol/L) Fig. 3 Effect of added P on Hg adsorption for the three shell samples. Symbols represent mean values of triplicates with all standard error below 15%.

No. 12 Table 3

Sample

a

b

c

Mercury removal using ground and calcined mussel shell Desorbed Hg for the three shell samples as a function of the added phosphate Initial Hg (µmol/L)

Desorbed Hg (µmol/kg)

1.25 12.5 25

7 (13) 18 (3) 37 (4)

1.25 12.5 25

4 (7) 58 (10) 85 (8)

1.25 12.5 25

7 (13) 45 (8) 71 (7)

Added phosphate (mmol/L) 0 0.16 8.07 16.15 0 0.16 8.07 16.15 0 0.16 8.07 16.15

Desorbed Hg (µmol/kg) 37 (4) 14 (0.8) 1 (0.0) 2 (0.0) 85 (8) 29 (1.7) 14 (0.9) 3 (0.2) 71 (7) 20 (1.2) 11 (0.7) 3 (0.2)

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Table 4 Fitting of kinetic data on Hg adsorption from the stirred flow chamber experiments referred to the three shell samples, with or without P addition Sample qmax (µmol/kg) a b c

6271 ± 56 399 ± 43 5194 ± 54

Sample qmax (µmol/kg) a b c

8797 ± 117 6843 ± 77 6695 ± 167

Without P ks (min−1 ) 0.015 ± 0.000 0.022 ± 0.001 0.017 ± 0.001 With P ks (min−1 ) 0.010 ± 0.000 0.013 ± 0.000 0.014 ± 0.001

R2 0.954 0.957 0.958 R2 0.932 0.940 0.871

Data are expressed as mean value of triplicates, initial Hg concentration is 50 µmol/L. Standard errors were below 15%. Desorbed Hg, expressed as percentages of that previously retained, is showed between brackets.

Data are expressed as mean value of triplicates, initial Hg concentration is 50 µmol/L. Standard errors were below 15%. Desorbed Hg, expressed as percentages that previously retained, is showed between brackets. The equation used was dq/dt = k(qmax -q)

decrease is higher in sample a, where Hg desorption reaches 1–2 µmol/kg when as low as 8 mmol/L P is added, with percentage desorption values being lower than 0.2%. Samples b and c show their lowest desorption values only when clearly higher P concentrations are added (16 mmol/L), then exhibiting 3 µmol/kg Hg desorption concentrations, representing 0.2% desorption.

6843 and 6695 µmol/kg (1.76, 1.37 and 1.34 mg/g) for samples a, b and c, respectively (Table 4). The increase in Hg adsorption in the presence of P was 40%, 71% and 29% for samples a, b and c, respectively; ks values were similar for all three shell samples, but lower to that obtained in the experiments performed in the absence of P, thus suggesting that the adsorption velocity is lower when P is present.

2.3 Stirred flow chamber experiment

2.3.2 Hg desorption

2.3.1 Hg adsorption Results from kinetic experiments with stirred flow chamber show that the highest Hg adsorption corresponds to sample a (calcined), followed by sample c (coarsely ground) and sample b (finely ground) (Fig. 4). The time needed to reach equilibrium was similar for all three shell samples (150–175 min). Data from Fig. 4 were fitted to Eq. (3), with R2 > 0.95. The Hg adsorption values in the equilibrium (qmax ) were 6271, 3998 and 5194 µmol/kg (1.26, 0.80 and 1.04 mg/g) for sample a (calcined shell), b (finely ground shell) and c (coarsely ground shell) respectively. However, the kinetic constant ks shows the sequence: sample b > sample c > sample a (Table 4); concretely, ks values were similar for all three shell samples, although slightly higher for sample b (0.022 min−1 ), which is the sample with the lowest amount of Hg adsorbed, thus suggesting that the adsorption velocity is inversely proportional to the adsorption capacity in the case of Hg. Figure 4 also shows the amounts of Hg adsorbed as a function of time in the presence of 3 mmol/L P. The addition of phosphate clearly increases Hg adsorption for all three shell samples, also shown by the fitting to adsorption data (Table 4); however, the kinetic constant ks slightly decreases for all three shell samples (Table 4). The presence of P favours Hg adsorption in all three shell samples, and the equilibrium takes longer to be achieved as compared to the results when no P was added. Data were also fitted to Eq. (3), and qmax values were: 8797,

Figure 5 shows similar Hg desorption results for all three shell samples. Data were fitted to Eq. (4). The amounts of Hg that could be desorbed under experimental conditions (q0 ) were 1373, 1338 and 1056 µmol/kg for samples a, b and c, respectively (Table 5), representing 21.9%, 33.5% and 20.3% of the amount of Hg previously absorbed by samples a, b and c, respectively. Sample b, the one with the highest desorption percentage, is also the shell sample with the lowest amount of Hg adsorbed. Desorption percentages are relatively low (given that stirred flow chamber enhances release phenomena (Pe˜na-Rodr´ıguez et al., 2010), suggesting that Hg adsorption is not easily reversible in the studied shell samples; kd values were similar to ks values, suggesting that adsorption and desorption velocities were Table 5 Fitting of kinetic data on Hg desorption from the stirred flow chamber experiments referred to the three shell samples, with or without P addition Sample qmax (µmol/kg) a b c

1373 ± 30 1338 ± 45 1056 ± 33

Sample qmax (µmol/kg) a b c

1948 ± 40 4092 ± 81 3232 ± 48

Without P kd (min−1 ) 0.024 ± 0.002 0.017 ± 0.001 0.023 ± 0.002 With P kd (min−1 ) 0.057 ± 0.005 0.018 ± 0.001 0.026 ± 0.001

The equation used was dqd /dt = kd (q0 –qd ).

R2 0.727 0.622 0.673 R2 0.781 0.816 0.845

Journal of Environmental Sciences 2013, 25(12) 2476–2486 / Susana Pe˜na-Rodr´ıguez et al.

0

100

9000 8000 Sample b 7000 6000 5000 4000 3000 2000 1000 0 0 100 9000 8000 Sample c 7000 6000 5000 4000 3000 2000 1000 0 0 100

500

100

200

300 Time (min)

400

500

500

9000 8000 7000 6000 5000 4000 3000 2000 1000 0 0

100

200 300 Time (min)

400

500

100

200

Adsorbed Hg (µmol/kg) 200 300 Time (min)

400

200 300 Time (min)

400

200 300 Time (min)

400

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9000 8000 a-P 7000 6000 5000 4000 3000 2000 1000 0 0

Sample a

Adsorbed Hg (µmol/kg)

9000 8000 7000 6000 5000 4000 3000 2000 1000 0

Adsorbed Hg (µmol/kg)

Adsorbed Hg (µmol/kg)

Adsorbed Hg (µmol/kg)

Adsorbed Hg (µmol/kg)

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500

9000 8000 7000 6000 5000 4000 3000 2000 1000 0

b-P

c-P

300 400 500 Time (min) Fig. 4 Hg adsorption kinetic experiments with stirred flow chamber. (a-P), (b-P), and (c-P) design experiments where 3 mmol/L P is added.

also similar. Figure 5 also shows the amount of Hg desorbed as a function of time in the presence of P. The desorption was higher for all three shell samples when P was present. Data were fitted to Eq. (4), and the values for q0 were 1948, 4092 and 3232 µmol/kg for samples a, b and c, respectively (Table 5). These values are 1.4 times higher than those in the absence of P for sample a, and 3 times higher than those in the absence of P for samples b and c. The amounts desorbed represent 22.1%, 59.8% and 48.5% of the Hg previously adsorbed by samples a, b and c, respectively. The high desorption percentages in samples b and c suggest that Hg adsorption is significantly reversible for these two not-calcined shell samples in the presence of P. However, for sample a the desorption percentage was similar to that found in the absence of P, indicating that adsorption has low reversibility in the case of the calcined shell. Further, kd values were higher than ks values, thus indicating that velocity was higher for desorption than for adsorption if phosphate is present in solution. In fact, when comparing kd values with those in the absence of P, samples b and c show similar desorption velocities,

0

while for sample a kd was much higher in the presence of P, indicating much more desorption velocity when P is present.

3 Discussion Batch experiments showed higher Hg(II) retention in calcined shell than in ground mussel shells. Pe˜na-Rodr´ıguez et al. (2010) signaled that higher concentrations of Fe and Al could provide a great number of Hg(II) binding sites in calcined mussel shells, the metal binding mechanism occurring via the OH− of oxyhydroxides; however, our results indicate that Fe, Al and also Mn concentrations are higher in the finely ground shells (Table 1), as previously found characterizing samples from the same factories (Pousada-Ferrad´as et al., 2010), which would not explicate the higher Hg(II) retention in calcined shell. Arnfalk et al. (1996), studying the retention of Hg and other elements on soils, remarked that Hg was the metal ion which interacted the least with the various soil types at pH > 4, which may be explained by formation of strong, soluble Hg(Cl)2 and Hg(Cl)2− 4 complexes, suppressing the adsorption onto

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soil matrixes; but in the present study, the highest total Cl concentration was found in coarsely ground shell (which also evidenced the presence of halite in X-ray diffraction), with the finely ground shell having the lowest concentration, thus these data would not explicate the higher Hg(II) retention on calcined mussel shell; contrary, Taboada et al. (2003), studying Cu(II) and Hg(II) retention on chitosan, indicated that, depending on Cl− concentration, mercury can form a great variety of complexes using chloride as bridge, and these complexes could produce an increment of retention, each site fixing more than one metal ion; however, this would not explicate the higher Hg(II) retention on the calcined shell in our study. Matlock et al. (2001) found irreversible precipitation of Hg by means of molecules having S in form of thiols; our data show that the highest S concentration corresponds to finely ground shell, with the calcined shell having the lowest concentration; thus, S concentration would not explicate the higher Hg(II) retention on calcined shell. Pe˜na-Rodr´ıguez et al. (2010) indicated that, taken into account that Ca and P are important components in calcined mussel shell, the general mechanism of Hg adsorption could be similar to that proposed by Hassan et al. (2008), finally giving POHg+ or CaOHg+ . In the present study, total Ca is clearly higher in calcined shell, while total P is higher in finely ground shell; Pousada et al. (2010) found the same trend studying shells from the same

factories, with water soluble concentrations being lower in calcined shell, thus suggesting precipitation of insoluble phosphates. The different Hg(II) retention values in our three kind of mussel shell could have some relation with surface area in the case of samples b and c, the higher the surface area the higher the Hg adsorption or KF value; but the overall relationship is not lineal, given that the highest KF value corresponds to sample a (calcined shell), with surface area lower than sample b (finely ground), thus lowering surface area importance; Cubillas et al. (2005a) indicated that BET surface area-normalized dissolution rates for calcite are of the same order of magnitude as those of aragonite, but geometric surface area-normalized calcite dissolution rates are 30% lower than corresponding aragonite rates; in our experiments, aragonite concentrations follows a sequence inverse to that of KF , with the lowest aragonite concentration in sample a, and the highest aragonite concentration in sample c. Sample a (calcined shell, the one with the highest Hg(II) retention capacity) had much more calcite (61%) than aragonite (7%), and this fact suggests that mechanisms focusing on aragonite, that have been previously used to explain the retention of certain divalent cations (other than Hg) on mussel shell or shell of other bivalves, could have scarce repercussion on Hg(II) retention; in this way, Prieto et al. (2003) indicated that Cd would experiment sorption onto aragonite by surface precipitation

Journal of Environmental Sciences 2013, 25(12) 2476–2486 / Susana Pe˜na-Rodr´ıguez et al.

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of (Cd,Ca)CO3 , this mechanism involving simultaneous dissolution-crystallization, while long-term uptake of Cd by calcite would occur by oriented overgrowth, armoring the substrate from further dissolution, and stopping the process when only a small amount of Cd has been removed from the fluid; in a similar way, Cubillas et al. (2005b) indicated that CdCO3 precipitates by epitaxial growth directly on calcite, efficiently slowing dissolution, but it will precipitate by random three dimensional heterogeneous nucleation on aragonite, leaving some pore space that allows dissolution to continue; also, K¨ohler et al. (2007) indicated that the uptake of Cd by biogenic aragonite takes place via heterogeneous nucleation of metal-bearing crystallites onto the shell surfaces. By the other side, Godelitsas et al. (2003) indicated that surface micro-topography of calcite changes significantly as a result of the interaction with Hg(II), with the initial rhombic etch pits induced by H2 O dissolution rapidly transforming to deeper etch pits, which would be due to Hg(II) sorption in the surface, and after ca. 2 hr these pits are interconnected to form larger ones, while Hg(II)-bearing phases grow onto the surface through a heterogeneous nucleation process, with the crystal growth of orthorhombic hydrated Hg(II) oxide on the surface of calcite being confirmed. Among our three kind of mussel shell, sample a (calcined) is the one with the highest calcite concentration, and it is also the only one containing dolomite; in this way, it is remarkable that Labidi (2008) found up to 87% Hg removal by means of crushed waste brick with mineralogical composition being essentially quartz, dolomite and calcite (hence without aragonite). P addition could increase Hg adsorption both in batch and stirred flow chamber experiments due to mercuryphosphate interactions (shell surface sorption, with homogeneous and/or heterogeneous precipitation), some of them being labile, but stable enough in the absence of non P-loaded convective flow. In this way, Oliva et al. (2011), studying the removal of Hg and other cations from water by means of a commercial apatite, indicated that apatite reacted with acid water by releasing phosphates that increased the pH up to 6.5–7.5, complexing and inducing metals to precipitate as metal phosphates, and concretely the removal of Hg could be associated with the formation of a solid phase consisting in Hg3 (PO4 )2 .

Contrary to that found in batch trials, stirred flow chamber experiments show that P addition increases Hg desorption, this effect being more pronounced in samples b and c. Note that, in the stirred flow chamber experiments, desorbed Hg is removed out of the chamber, which could facilitate further Hg desorption when comparing with batch type experiments. In the desorption experiments performed with the stirred flow chamber, pH values are lower when P is added than when P is absent (Fig. 6). However, within the pH range at which Hg desorption increase occurs (pH between 7 and 9), lower pH values could cause lower Hg desorption if the behaviour would be comparable to that reported by Jing et al. (2007) for soils, so that it would not contribute to explain the higher Hg desorption in stirred flow chamber when P is present. In our experiments, the presence of P would increase Hg desorption in the stirred flow chamber due to the previous formation of labile phosphate-mercury bindings, followed by the desorption phase establishing the non Ploaded convective flow, which would facilitate the rupture of those labile phosphate-mercury bindings, then favouring Hg diffusion out of the chamber. On the contrary, the presence of P would decrease Hg desorption in the batch type experiments due to the absence of convective flow, which would minimize rupture of bindings and further diffusion of Hg and P. Previous works have reviewed the use of various lowcost sorbents to remove Hg from water (for example: Babel and Kurniawan, 2003; Bailey et al., 1999), but all of them show advantages and disadvantages, and need detailed cost-benefit considerations (Miretzky and Fernandez-Cirelli, 2009). Mussel shells are waste materials needing to be productively recycled, and one possible use could be the treatment of polluted waters, including Hg removal. Our results on Hg maximum retention capacity (qmax ) are above some corresponding to materials such as char, pozzolana and yellow tuff reported by Miretzky and Fernandez-Cirelli (2009), even if compared with our stirred flow experiments, where reaction time is much lesser than in classical batch adsorption trials, although they are below qmax corresponding to many other low-cost materials. Globally, taking into account our results, the fact that the presence of phosphate in solution did not cause a decrease With P

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in Hg retention on mussel shells, but increased it, together with the low degree of Hg desorption experienced (specially in calcined shell, and unless when convective flow takes place), indicating that Hg adsorption is not easily reversible in this mussel waste, would be of importance to facilitate the use and recycling of these materials in the treatment of polluted waters. Future research could be directed to elucidate the efficacy of mussel shell to remove pollutants from waters where several heavy metals and other cationic and anionic contaminants (in addition to Hg and phosphate) would be present simultaneously, and to deeply understand some mechanisms contributing to explicate the retention and release processes affecting the investigated pollutants.

4 Conclusions The calcined mussel shell is the material with the highest potential Hg adsorption and lowest Hg desorption among the three kind of mussel shell studied. This shell has the highest calcite percentage and is the only one containing dolomite. The presence of phosphate increases Hg adsorption for all three shell samples, both in batch type and stirred flow chamber experiments. The P addition also decreases Hg desorption in batch type experiments, but increases desorption of Hg in the stirred flow chamber experiments. Both calcined and ground mussel shell would be useful for Hg removal, with or without P in solution, with calcined shell presenting better results but also higher cost of production from the raw waste material. Acknowledgments This work was funded by the INCITE program of the Galician Council of Innovation and Industry (Spain) (Ref. 08PXIB383190PR). Bermudez-Couso A. is funded by the pre-doctoral programme from the University of Vigo (Spain). Abonomar S. L. (Spain), and Calizamar S. L. (Spain) are acknowledged for kindly supplying mussel shell batches.

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