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A forward osmosis membrane system for the post-treatment of MBR-treated landfill leachate
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A forward osmosis membrane system for the post-treatment of MBR-treated landfill leachate Ying Dong, Zhiwei Wang, Chaowei Zhu, Qiaoying Wang, Jixu Tang, Zhichao Wu
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S0376-7388(14)00639-5 http://dx.doi.org/10.1016/j.memsci.2014.08.023 MEMSCI13135
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Journal of Membrane Science
Received date: 28 May 2014 Revised date: 12 July 2014 Accepted date: 11 August 2014 Cite this article as: Ying Dong, Zhiwei Wang, Chaowei Zhu, Qiaoying Wang, Jixu Tang, Zhichao Wu, A forward osmosis membrane system for the posttreatment of MBR-treated landfill leachate, Journal of Membrane Science, http: //dx.doi.org/10.1016/j.memsci.2014.08.023 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting galley proof before it is published in its final citable form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
A forward osmosis membrane system for the post-treatment of MBR-treated landfill leachate Ying Dong1, Zhiwei Wang1,*, Chaowei Zhu2, Qiaoying Wang1, Jixu Tang1, Zhichao Wu1 1
State Key Laboratory of Pollution Control and Resource Reuse, School of Environmental
Science and Engineering, Tongji University, Siping Road 1239, Shanghai 200092, PR China 2
Chinese Research Academy of Environmental Sciences, Beijing 100012, PR China
Revised manuscript submitted to Journal of Membrane Science for possible publication
*Corresponding
Author.
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+86-21-65980400
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Abstract Filtration behaviours, membrane fouling and cleaning were investigated in a forward osmosis (FO) membrane system used to post-treat the effluent of a membrane bioreactor (MBR) fed with landfill leachate (LFL). In short-term tests, it was observed that the water flux with the membrane active layer facing the draw solution (AL-DS mode) was lower than that with the membrane active layer facing the feed solution (AL-FS mode) for LFL treatment. Mathematical models could well simulate the flux evolution of AL-FS mode while the flux of AL-DS deviated from the modeling curve, suggesting that fouling could be rapidly developed within 1 h filtration for AL-DS mode. During long-term filtration, about 98.6% of COD, 96.6% of TP, and 76.9% of ammonium were rejected by the FO system. A decrease of water flux was also observed with an increase in operation time. Confocal laser scanning microscopy and Fourier transform infrared spectroscopy confirmed the existence of polysaccharides and proteins in the fouling layer. Inorganic fouling was mainly caused by Ca, Na, Mg, K, Si, Fe and Al. It was also found that the effect of cake enhanced concentration polarization played an important role during long-term operation. About 88.9% of the permeate flux was recovered after hydraulic cleaning while it reached 98.9% of the initial flux after chemical cleaning, indicating that chemical cleaning was needed to eliminate irreversible fouling and to recover membrane permeability during long-term operation. Keywords: Forward osmosis; landfill leachate; membrane fouling; membrane cleaning; wastewater treatment
1. Introduction Landfilling is still the most widely used method for municipal solid waste (MSW) disposal throughout the world, though recycling, composting and incineration are currently being encouraged [1, 2]. One of major concerns in landfills is the generation of leachate by the waste within the system and/or the infiltration of groundwater/rainfall. Landfill leachate (LFL) usually has high concentrations of organic pollutants, toxic materials (xenobiotic organics), ammonia, heavy metals and inorganic salts, and its characteristics can be significantly changed due to the variations of MSW composition, age, climatic conditions, landfill design and operational practice 2
[3]. If it is not properly collected and treated, it can potentially contaminate nearby surface and ground water. Development of efficient LFL treatment methods plays an important role in solving the potential threats of MSW landfills. To date, a number of treatment protocols have been proposed, e.g., physico-chemical methods including adsorption, stripping, coagulation, oxidation and membrane separation, and biological processes [4, 5]. For achieving high treatment efficiencies, physico-chemical treatment is always combined with biological processes. Physico-chemical units are placed either as pretreatment to reduce the loading rate for biological processes or as post-treatment to reach a high quality discharge standard. With the rapid development of membrane technology, combination of membrane bioreactor (MBR) with nano-filtration (NF) or reverse osmosis (RO) has become increasingly popular in LFL treatment [4, 6-8]. However, efforts should be taken to mitigate membrane fouling and reduce the energy consumption of NF/RO processes. Forward osmosis (FO) is an osmotically-driven membrane process, in which the driving force for separation is the difference in chemical potential between a concentrated draw solution and a broad range of aqueous solutions [9]. A semi-permeable FO membrane can retain solutes but allow water to transfer through the membrane. Compared to NF/RO, FO has several advantages such as no need for hydraulic pressure, high rejection of pollutants, and low membrane fouling propensity. In the past decade, FO has been used for oil and gas wastewater, and industrial and municipal wastewater treatment [10]. There is also a growing interest in the application of FO to LFL treatment [11]. The use of FO to replace NF/RO might be promising in addressing the existing obstacles. However, detailed information on the performance of FO in LFL treatment is scarce. In this study, an FO membrane system was used to post-treat the effluent of an MBR fed with landfill leachate. Batch experiments and long-term filtration behaviours were monitored in the FO membrane system. Membrane fouling and cleaning were also investigated. The obtained results are expected to provide a sound understanding of FO membrane performance in treating LFL.
3
2. Materials and methods 2.1. Experimental set-up A lab-scale FO system, as shown in Fig. 1, was used in this study. The membrane cell comprises equally structured rectangular channels on both sides of the membrane. The dimensions of each channel are 100 mm long, 20 mm wide and 1.5 mm deep. The effective membrane area is approximately 0.002 m2. Two variable-speed diaphragm pumps (Mini-pump, Benquan, Shanghai, China) were used to recirculate the feed and draw solutions, respectively. A cross flow velocity of 25 cm/s was maintained on both sides of the membrane through recirculation. The water flux was determined by measuring the weight changes of the feed solution in real time with an electronic balance (Ohaus, New Jersey, USA) connected to a laptop and a data logging system. All experiments were carried out at room temperature (25 ± 0.5 oC). Before use, the membrane was soaked in DI water for 24 hours and further washed with DI water under a cross-flow velocity 25 cm/s for 30 minutes. The volumes of the feed and draw solution tanks are both 6 L. The FO membrane used in this study was provided by Hydration Technologies Innovation (HTI, Albany, USA). It is made of cellulose triacetate (CTA) with embedded polyester mesh as mechanical support.
Fig. 1. Schematic of the lab-scale FO system. 2.2. Operating conditions 2.2.1. Short-term filtration Since the membrane has an asymmetric structure, the water flux of both membrane orientations (i.e., AL-FS with the membrane active layer facing the feed solution, and AL-DS with the membrane active layer facing the draw solution) was determined. Firstly, DI water was used as 4
feed solution to obtain the DI water flux using different concentrations of draw solutions. Then, the effluent from an MBR fed with LFL was used as feed solution. The characteristics of the feed solution are shown in Table 1, and the detailed information on this MBR can be found elsewhere [12]. Each batch test lasted for 1 hour. After each test, the feed and draw solutions were changed to DI water, and the membrane was cleaned for 30 min under the cross-flow velocity of 25 cm/s. Analytical grade sodium chloride (NaCl) was used to prepare draw solutions with concentrations ranging from 0.5 M to 4.0 M.
Table 1 Characteristics of the feed solution (n= 3). Items
Concentration (mg/L except for osmotic pressure)
COD
696±20
TOC
215±10
TDS
7100±47
TN
143±12
TP
0.3±0.0
2+
925±22
Mg2+
324±14
Cl-
3676±58
Osmotic pressure (mOsm/kg)
785±4
Ca
2.2.2. Long-term filtration During the long-term filtration experiment, 3.0 M NaCl was used as the draw solution and the MBR-treated effluent was adopted as the feed solution. The feed and draw solutions were replaced with fresh ones every 24 hours in order to avoid significant changes in their concentrations. The AL-FS orientation was used in order to avoid dramatic flux decline. No cleaning was performed until the water flux was decreased to about 60% of the initial. Hydraulic cleaning was carried out for 1 h under the cross-flow velocity of 25 cm/s by changing the feed and draw solutions to DI water. After 5 runs, the membrane was taken out from the cell, and 1-h chemical cleaning under the cross flow velocity of 25 cm/s was conducted using 1% Alconox (White Plains, USA) solution as the feed and draw solution. Afterwards, the feed and draw solution was changed to DI water for 5
10 min under cross flow in order to remove the remaining Alconox solution. Another run was performed after the chemical cleaning to observe the flux recovery using the MBR-treated effluent as feed solution. A parallel experiment was carried out in order to obtain the fouled membrane samples for fouling characterization as mentioned in Section 2.4. 2.3. Modeling FO performance An analytical model, incorporating the effect of internal concentration polarization (ICP), has been developed by Loeb et al. [13] for evaluating FO performance. According to this model, the FO water flux (Jv) can be worked out by Eq. (1) for the AL-FS orientation and Eq. (2) for the AL-DS orientation [14]. ⎛ Aπ draw + B ⎞ J v = K m ln ⎜ ⎟ ⎝ Aπ feed + J v + B ⎠
(1)
⎛ Aπ − Jv + B ⎞ J v = K m ln ⎜ draw ⎟ A π feed + B ⎝ ⎠
(2)
In Eqs. (1) and (2), πdraw and πfeed are the osmotic pressure of the draw solution and feed solution, respectively. A and B are the water and salt permeability of the rejection layer, respectively, which can be determined by RO filtration tests as described by Tiraferri et al. [15]. Km, the mass transfer coefficient, is related to the ICP phenomenon within the porous support layer. Km can be worked out using the solute diffusion coefficient Ddraw divided by the membrane structure parameter Sme.
Km =
Ddraw Ddraw ⋅ ε me = S me tme ⋅ τ me
(3)
where εme, tme and τme are the porosity, thickness and tortuosity of the membrane support layer, respectively. The diffusion coefficient (D) can be worked out through Eq. (3) using Km obtained from Eq. (1) and Eq. (2) and the membrane structure parameter S=3.93×10-4 m [16]. Eqs. (1) and (2) are valid for well-defined feed and draw solutions in most cases [17, 18]; however, the effects of fouling under real operating conditions are not explicitly accounted for. To address this issue, Lay et al. [19] proposed a fouling-incorporated water flux model for an FO system based on a resistance-in-series approach for a fouling condition with cake enhanced concentration polarization (CECP). The water flux of an FO system with the AL-FS orientation 6
can be expressed by Eq. (4).
B⎞ ⎡⎛ J v = A ⋅ ⎢⎜ π draw + ⎟ ⋅ e −( J v A⎠ ⎝ ⎣
Km )
B⎞ ⎛ − ⎜ π feed + ⎟ ⋅ e( J v A⎠ ⎝
kCECP )
⎤ (4) ⎥ ⎦
where A and B are the overall water and salt permeability coefficients, which are related to the coefficients of a membrane (subscript ‘me’) and fouling layer (subscript ‘la’) as follows [19].
1 1 1 = + A Ame Ala
(5)
1 1 1 = + B Bme Bla
(6)
The coefficient kCECP can be expressed via a mass transfer coefficient in Eq. (7) [19].
kCECP =
Dml ⋅ ε la δ la ⋅ τ la
(7)
where Dml is the diffusion coefficient on average of the solutes within the fouling layer, and εla, δla, and τla are the porosity, thickness and tortuosity of the fouling layer, respectively. A smaller kCECP indicates that the CECP effects are more significant. If kCECP=∞, the CECP effects are negligible. The above-mentioned Eq. (4) can be rewritten as Eq. (1) by setting kCECP=∞. For AL-DS orientation in an FO system, the water flux can be modeled using Eq. (8) under the assumption that the external concentration effects are negligible.
B⎞ ⎛ B⎞ ⎡⎛ J v = A ⋅ ⎢⎜ π draw + ⎟ − ⎜ π feed + ⎟ ⋅ e( J v A A⎠ ⎝ ⎠ ⎝ ⎣
K m+foul )
⎤ (8) ⎥ ⎦
where the parameter Km+foul is a complex overall mass transfer coefficient that couples both ICP and fouling mechanisms. In the absence of fouling or other performance diminishing effects, Km+foul is equal to Km [20], and Eq. (8) can be reduced back to Eq. (2). In this study, Eqs. (1) and (2) were used to model the water flux under AL-FS and AL-DS orientations respectively in the batch experiments. Eq. (4) and Eq. (8) were also employed to evaluate the development of fouling in the bath filtration for LFL treatment. For the long-term experiments under AL-FS orientation, Eq. (4) was adopted to assess the filtration behaviour.
7
2.4. Analytical methods 2.4.1. Morphological analyses For scanning electron microscopy (SEM) analysis, a scanning electron microscope (Model XL 30, Philips, Netherlands) equipped with an energy-dispersive X-ray spectroscope (EDX) was used. Before observation, all samples were dried in vacuum at room temperature (25 °C), and sputter-coated with a thin layer of gold. The fouled FO membrane was also observed. Elemental analysis of the foulant on the membrane surface was carried out using EDX immediately after SEM observation. A confocal laser scanning microscope (CLSM) (Nikon A1, Tokyo, Japan) was used in both fluorescent mode and reflective mode to visualize the membranes. The samples can be stained using the methods described by Wang et al. [21] and Chen et al. [22]. Lateral x-y scans were performed with a ×20 objective, and images were stored with a resolution of 512 × 512 pixels (representing an area of 360 × 360 μm). The scans were conducted from the active layer (active layer facing the objective lens). Images were collected at a series of sampling depths with a scanning depth increment of 1.65 μm in the z-direction to obtain the internal structure of membrane samples. For this study, we set z = 0 at the membrane active layer surface, with the positive z-direction pointing toward the supporting layer [21]. SEM and CLSM images were analyzed using image analysis software (ImageJ, NIH, USA) to obtain the information of pore size and porosity of the virgin membrane samples. The CLSM and SEM images were converted into 8-bit gray scale images, which were then transformed to black and white images by adjusting the threshold value. The porosity was determined as the ratio of the total void area to the gross area analyzed using the software. The sampled area for each analysis was about 100 × 100 μm for CLSM images, and about 70 × 88 μm for SEM images. In both CLSM and SEM, at least 3 images of each membrane were determined [21] . 2.4.2. Fourier transform infrared spectroscopy (FTIR) For FTIR studies, fouled membrane collected from the long-term experiment was pretreated in a vacuum freeze drying machine for 24 hours. Then the dried membrane together with the foulant was directly analyzed by FTIR spectroscopy with an OMNI sampler and an attenuated total 8
reflectance (ATR) accessory (Nicolet 5700, Thermo Electron Corporation, USA), in order to determine the functional groups of the foulants. The FTIR of the clean membrane was used as the blank, and its transmittance was subtracted from the spectra of the fouled membrane. The FTIR spectra were analyzed based on the Sadtler handbook of Infrared Spectra. 2.4.3. Wastewater characteristics Analyses of chemical oxygen demand (COD), total nitrogen (TN), and total phosphate (TP) were performed based on the standard Chinese NEPA methods [23]. The concentrations of Ca and Mg
were
determined
using
an
inductively
coupled
plasma
emission
spectrometer
(ICP-Agilent720ES, Agilent Technologies, Santa Clara, USA). Total dissolved solids (TDS) were measured using a DDSJ-308A conductivity meter (Leici, Shanghai, China). Total organic carbon (TOC) was determined by using a VCPN total organic carbon analyzer (Shimadzu Corporation, Kyoto, Japan). The measurement of osmotic pressure was carried out by using an FM-9J osmometer (Yida, Shanghai, China). The rejection of pollutants in the FO system is obtained from Eq. (9). R=
VC ×100% V0C0
(9)
In Eq. (9), V0 and V are the initial volume of the feed solution and the final volume of the feed solution prior to the change into fresh one every day, respectively. C0 and C are the initial and final concentrations of the pollutants measured in the feed solution, respectively. The rejection values of pollutants are reported as the average values and standard deviations during 36-d operation.
3. Results and discussion 3.1. Osmosis membrane characterization and short-term filtration The characteristics of the virgin FO membrane are summarized in Table 2. The FO membrane, as shown in Fig. S1 (see the supporting information), has a dense active layer (9 μm) and a support layer (39 μm). The polyester (PE) mesh fibers at a spacing of about 110 μm could be clearly seen from the cross section and the support layer of the membrane. A larger fiber-to-fiber distance of about 160 μm has been also reported for other membranes [21]. The total thickness of 9
the FO membrane is 48 μm, much thinner than standard RO membranes in order to enhance FO membrane permeability [9]. The membrane’s support layer (Fig. S1) has a mean pore size of 5.2 μm. The relatively larger pore size can mitigate internal concentration polarization phenomenon. It has been also reported that the improvement of hydrophilicity can facilitate the FO membrane permeability [9]. However, the AL of the virgin membrane used in this study is hydrophobic, indicating there is much space to further modify the membrane. The water permeability (A) and salt permeability (B) of the membrane are consistent with the results of previous publications [20, 24, 25]. Table 2 Virgin membrane characterization. Parameters
Value
Thickness of AL (μm)
9± 2
Thickness of SL (μm)
39± 6
Mean pore size of SL (μm)
5.2±0.9
Contact angle of AL (°)
90.6 ± 2.1
Contact angle of SL (°)
70.4 ± 0.7
Water permeability (A) (L/(m2 h bar))
1.33±0.12
Salt permeability (B) (L/(m2 h))
1.46±0.15
Salt rejection rate (%)
92.3±0.3
Note: AL: active layer; SL: support layer. The values are given as average ± standard deviation. The mean pore size of SL was obtained from the SEM image. Number of measurements (n): n=20 for thickness; n=100 for pore size; n=6 for contact angle; n=3 for the rest of the parameters.
In order to further examine the inner structure of the membrane, layer-by-layer CLSM scanning was carried out. The CLSM images and corresponding pore size and porosity of the inner structure are illustrated in Fig. 2. There are no visible pores in the active layer (z=0 μm), which might be due to the dense property of the active layer and the low magnification of CLSM. Smaller pores (z<10 μm) were found near the active layer, and larger pores (z>10 μm) were observed on the support layer side, which is consistent with the results reported by Wang et al. [21]. It is interesting to find that the inner structure of the SL has a larger pore size and porosity than the surface of the SL (z=49 μm). This may increase the difficulty to eliminate ICP through increasing cross-flow velocity on the SL side. 10
(a) (c) (b) 100 μm 100 μm 100 μm (f) (d) (e) 16
100 μm
100 μm
Pore diameter (μm)
14
Pore diameter Porosity
12 10
100 90 80 70 60 50 40 30 20 10 0
Porosity (%)
8 6 4 2 0 0 10 20 30 40 50 Distance from the AL surface (μm)
Fig. 2. CLSM images of a clean FO membrane viewed from the active layer side. (a) z=0 μm; (b) z=8.25 μm; (c) z=16.5 μm; (d) z=24.75 μm; (e) z=33 μm; and (f) the pore size and porosity of the membrane inner structure as a function of z (the distance from the membrane active layer surface).
Table 3 summarizes the water flux, mass transfer coefficient and solute diffusion coefficient of the virgin membrane using DI water as the feed solution. A strong correlation between the water flux and draw solution concentration was observed, which was due to the increasing osmotic pressure difference across the membrane. However, the flux was not evidently changed at higher NaCl solution concentration, suggesting that the water flux is non-linearly correlated with the draw solution concentration as a result of ICP [17]. The effect of ICP is supposed to be much severer at higher water fluxes [20, 24]. The membrane orientation has a significant effect on the water flux. The water flux of DI water in AL-DS mode is much higher than that in AL-FS mode with the same draw solution concentration.
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Table 3 Water flux (Jv), mass transfer coefficient (Km) and solute diffusion coefficient (D) of the membrane using DI water as the feed solution. NaCl
Osmotic
concentration
pressure
(M)
(πdraw, bar)
(L/(m h))
(L/(m h))
0.5
25
9.8±0.2
8.71±0.32
1.0
50
13.7±0.1
1.5
75
2.0
AL-FS Jv 2
AL-DS
Km 2
D -10
Jv 2
(10 m /s)
Km 2
2
D -10
(10 m2/s)
(L/(m h)
(L/(m h)
9.51±0.34
14.2±0.2
5.37±0.10
5.87±0.10
9.13±0.11
9.97±0.12
26.0±0.3
7.74±0.11
8.45±0.12
19.5±0.1
12.38±0.12
13.52±0.13
30.2±0.2
7.77±0.06
8.49±0.06
100
21.9±0.2
12.51±0.13
13.66±0.14
39.2±0.2
9.38±0.04
10.24±0.04
2.5
125
24.4±0.1
13.05±0.08
14.25±0.09
44.4±0.2
10.01±0.05
10.93±0.05
3.0
150
26.9±0.2
13.74±0.12
15.00±0.13
48.4±0.3
10.41±0.07
11.36±0.08
4.0
200
28.0±0.2
12.69±0.13
13.86±0.14
51.8±0.3
10.37±0.06
11.32±0.07
Note: n=3 for Jv, Km and D; πfeed=0 bar, A=1.33 L/ (m2 h bar), B=1.46 L/ (m2 h), S=3.93×10-4 m for calculating Km and D using Eqs. (1)~(3).
In order to further compare the water flux of FO membranes using DI water and LFL as the feed solutions, Fig. 3 is drawn to illustrate the obtained experimental flux and the modeled flux as functions of osmotic pressure difference. It is evident that the water flux of FO membranes deviates from linear curve (Jv=A(πdraw-πfeed)) based on the classical solution-diffusion theory [9], indicating that the fouling (e.g., ICP using DI as feed solution and complex fouling due to the use of LFL as feed solution) can significantly impact the water fluxes in real applications [13]. It was also found that the mathematic models, as expressed by Eqs. (1) and (2), could well fit the experimental data in both AL-FS and AL-DS modes using DI water as feed solution. The flux deviation from the linear correlation can be well explained by dilutive ICP in the SL of AL-FS mode and concentrative ICP in AL-DS mode. It has been reported that A has greater impacts on water flux than Km with smaller osmotic driving force (e.g., (πdraw-πfeed) ≤10 bar); however, Km dominates the influence with greater driving force ((πdraw-πfeed) ≥20 bar) [20]. The membrane used in this study has a higher Km value and a similar A compared to the membranes in the reference [20], suggesting that higher membrane fluxes could be achieved in particular in the range of larger driving force. 12
For LFL, it is observed that the water flux in the AL-FS mode is higher than that in the AL-DS mode during the short-term filtration (1 h). Interestingly, this is contrary to the results of using activated sludge as the feed solution [26]. To elucidate the mechanisms, we used Eqs. (7) and (8) to simulate the water fluxes of LFL using AL-FS and AL-DS modes, respectively. It can be seen that the experimental data for AL-FS mode is well modeled by Eq. (7) by setting kCECP= ∞ and
πfeed= 20.1 bar (see Fig. 3). An infinite (∞) value of kCECP indicates that the CECP effect is negligible [20]. The results suggested that the CECP was not obviously formed in the AL-FS mode for LFL treatment in short-term filtration. However, it was interestingly found that the experimental fluxes were much lower than the modeled values under the AL-DS mode for LFL treatment by using Eq. (8) (setting Km+foul= Km and πfeed= 20.1 bar), demonstrating that the parameter Km+foul could not be simplified as Km. The parameter Km+foul is considered as a complex overall mass transfer coefficient that incorporates both ICP and fouling processes in the AL-DS mode [20]. Take the DS concentration of 3.0 M as an example, the Km+foul for LFL calculated from Eq. (8) is 5.33 L/(m2 h), while the value of Km for DI water is 10.41 L/ (m2 h). This indicates that the fouling of FO membranes in the AL-DS mode can be rapidly developed within 1 h filtration. Therefore, for real wastewater treatment, the AL-FS mode is always adopted in long-term filtration. Compared with the water flux of using activated sludge as the feed solution under the same experimental conditions (DS: 0.5 M NaCl) [26], the water flux of LFL was much lower. It demonstrates that the fouling conditions of LFL are more critical, which can be attributed to the more complicated composition and the higher salts in LFL.
13
AL-DS DI
AL-FS DI
AL-FS LFL AL-DS LFL
Fig. 3. Modeled and experimental water flux as a function of osmotic pressure difference. symbols: □ experimental data of water flux using AL-DS mode and feed solution DI water; ■ experimental data of water flux using AL-FS mode and feed solution DI water;● experimental data of water flux using AL-FS mode and feed solution (LFL); ○ experimental data of water flux using AL-DS mode and feed solution LFL. In calculating the modeled water flux, πfeed=0 bar for DI water, πfeed=20.1 bar for landfill leachate (LFL), A=1.33 L/ (m2 h bar), and B= 1.46 L/ (m2 h); the values of Km were obtained by fitting the measured data (Km=2.513ln(Δπ)+0.5475, R2=0.8076 in AL-FS mode; Km=2.5516ln(Δπ)−2.6481, R2=0.9481 in AL-DS mode). The red solid line is the modeled flux using Jv=A(πdraw-πfeed), and other lines indicate the modeled flux using Eqs. (1) and (2) based on the adopted membrane orientation.
3.2. Membrane separation performance during long-term operation Fig. 4 shows the water flux evolution of LFL in AL-FS mode during 36-day operation using the AL-FS mode. The initial flux at the beginning of the experiment was 18.0 L/ (m2 h), which is in agreement with the experimental water flux as shown in Fig. 3 during the short-term filtration. Since both the feed and draw solutions were replaced with fresh ones every day in this study, the impact of concentration changes on water flux was limited. However, distinct flux decline during each run can be observed, indicating that fouling was a critical issue in the long term FO 14
performance [14]. Hydraulic cleaning was carried out in the first 4 runs (indicated by the blue circles), while chemical cleaning (red circle) was performed at the end of fifth run. It is also evident in Fig. 4 that irreversible fouling is gradually developed during the long-term filtration as revealed by the dashed line, indicating that chemical cleaning should be periodically carried out in order to sustain the operation.
Fig. 4. Water flux during the long-term filtration experiment under the AL-FS mode for LFL treatment.
Fig. 5 shows the average retention efficiency of the pollutants in the LFL. About 98.6% of COD, 96.6% of TP, and 76.9% of ammonium were rejected by the FO system. The rejection of Ca2+ and Mg2+ was also very effective. The lower rejection of ammonium compared with Ca2+ and Mg2+ should be attributed to its smaller ionic radius or hydrated ionic radius. In addition, Ca2+ and Mg2+ could also result in the scaling on the membrane surface [16, 27].
15
Fig. 5. The rejection of pollutants during the long-term filtration experiment (n=36). 3.3. Membrane fouling and cleaning In order to examine the fouling behaviours, the fouled membrane in the long-term filtration experiment was taken out from the membrane cell after the fifth fouling test run without any cleaning and analyzed by SEM and CLSM. Fig. 6 (a) and Fig. 6 (b) show the SEM images of the active and support layer of the fouled membrane, respectively. As shown in Fig. 6 (a), an obvious deposition of foulants on the active layer was observed, while the fouling was not severe on the support layer (Fig. 6 (b)). Unlike a cake layer formed on membrane surfaces in an FO-MBR [14], a number of crystallization could be found on the active layer, indicating that the dominant foulants are different in various FO systems. The EDX results (Table 4) show that Ca, Na, Mg, K, Si, Fe and Al are the main inorganic elements in the foulants, especially calcium, which is considered to have negative effects on the membrane performance [28]. (a)
(b)
200 μm
200 μm
Fig. 6. SEM images of fouled FO membrane surfaces (a) SEM image of active layer; (b) SEM image of support layer.
16
The components of the foulants on the active and support layers were also analyzed using FTIR. Fig. 7 shows the FTIR spectra of the fouling layer on both sides. The broad peak at 1045 cm−1, 1232 cm−1 and 1742 cm−1 are associated with polysaccharides or polysaccharide-like substances. The peak at 1233 cm−1 is assigned to the presence of sulfate ester groups which is a characteristic component in fucoidan and sulphated polysaccharides, while the peak at 1742 cm−1 is assigned to the carboxylic acid ester form (C=O) in alginic acid standard [29]. The broad peak at 1045 cm−1 is also related to polysaccharides or polysaccharide-like substances [30]. Signals at 1515 cm−1 correspond to the amide II bands [29], indicating the presence of proteins. The FTIR results indicate that protein and polysaccharide substances may be the major organic foulants.
Table 4 EDX microanalysis of fouled FO membrane. Weight percentage (%)
O
F
Na
Mg
Al
Si
P
Cl
K
Ca
Fe
Active layer
53.21
4.59
6.84
3.59
0.21
2.01
0.09
8.08
2.16
18.83
0.39
Support layer
62.86
--
15.56
--
--
--
--
19.76
--
1.82
--
Note: -- not detectable.
Fig. 7. FTIR spectra of the membrane foulants in the long-term fouling experiment
For further examining the fouling layer on the active layer surface of the FO membrane, fluorescent staining and CLSM observation were performed. Figs. 8 (a)-(c) demonstrate the distributions of α-polysaccharides (Con A), β-polysaccharides (Calcofluor white) and proteins (FITC). According to the fluorescent intensity presented in Fig. 8, abundant proteins and polysaccharides (especially α-polysaccharides) exist in the fouling layer. They have been reported 17
to be the major components of organic fouling during membrane filtration of water and wastewater [31-33]. In combination with EDX analysis, the fouling could be interpreted as the comprehensive impacts of organic foulants, inorganic precipitates and/or organic-inorganic complexes. The formation of inorganic scaling might be ascribed to the severe effects of concentration polarization caused by inorganic ions existing in the LFL and the high rejection property of the selective layer of the FO membrane. Furthermore, the organic fouling layer formed on the membrane surface may have hindered the diffusion of inorganic foulants back to the feed solution, promoting the inorganic precipitation. It is suggested that synergistic effects between organic and inorganic foulants can result in much severer fouling of FO membranes [34]. (a)
(b)
100 μm
100 μm
(c)
(d)
100 μm
100 μm
Fig. 8. The CLSM images of stained fouling layer (a) CLSM image of α-polysaccharides (Con A); (b) CLSM image of β-polysaccharides (Calcofluor white); (c) CLSM image of proteins (FITC); (d) combination of the individual images.
Table 5 summarizes the membrane permeability of fouled and cleaned membrane during the long-term filtration. A dramatic decline of A and B of the fouled membrane was observed compared with the virgin membrane. Hydraulic cleaning led to an increase of A and B, and chemical cleaning further recovered the value almost back to the initial. This suggests that the effects of organic foulants could not be ignored in this study except for inorganic fouling [14]. Ala 18
and Bla can be worked out through Eqs. (5) and (6), respectively, and their values in Table 5 clearly show that the decrease of A and B is associated with the decrease of Ala and Bla from the infinite to a certain value due to fouling. Higher values of Ala and Bla can result in larger A and B. The parameter kCECP can indicate the fouling condition; a lower value denotes a severer fouling. It clearly shows the CECP plays an important role in the fouling process during the long-term operation. Regarding the water flux recovery, about 88.9% of the permeate flux was recovered after the hydraulic cleaning while the value reached 98.9% of the initial after the chemical cleaning. This indicates that chemical cleaning is needed to address irreversible membrane fouling and to recover membrane permeability during long-term operation.
Table 5 Membrane permeability of the fouled membrane and the cleaned membrane (hydraulic and chemical cleaning). B
A
Ala 2
Bla 2
Jv
(L/(m h
(L/(m
(L/(m h
(L/(m
(L/(m2
bar))
h)
bar))
h))
h))
Virgin membrane
1.33
1.46
∞
∞
Fouled membrane
0.72
0.85
1.57
0.80
0.92
1.25
1.37
Membrane samples
2
2
kCECP 2
Flux recovery
(L/(m h))
(%)
18.0
∞
-
2.03
8.2
6.75
-
2.01
2.49
16.0
65.59
88.9
20.78
22.22
17.8
72.73
98.9
Membrane after hydraulic cleaning a Membrane after chemical cleaning a
Τhe value is for hydraulic cleaning performed at the end of the fourth run. πdraw=150 bar, πfeed=20.1 bar,
Km=13.74 L/(m2 h) for calculating kCECP by Eq. (4).
It has to be pointed out that daily replacement of both draw and feed solutions was performed in order to avoid the decline of the driving force (πdraw –πfeed) in this study. Under this circumstance, the water recovery rate is very low. For practical applications, the water recovery rate needs reaching a high value, and also continuous flow mode should be adopted. For instance, the continuous inflow of the influent could ensure that the concentration of the feed solution is maintained at relatively constant value. The draw solution can be also kept constant through adding more concentrated salt solutions or concentrating draw solution using other devices (e.g., 19
reverse osmosis). Therefore, the obtained results in this study can reflect some practical cases that the concentration of the feed solution is maintained at about 700 mg/L in terms of COD. Further studies are needed in order to simulate more practical conditions of using FO to treat landfill leachate.
20
4. Conclusions An FO membrane system was used for post-treatment of landfill leachate from a membrane bioreactor, and fouling behaviours, membrane fouling and cleaning were studied. In short-term batch experiments, the water flux in the AL-FS mode was higher than that in the AL-DS mode, indicating that the fouling of FO membranes in the AL-DS mode was rapidly developed within 1 h filtration. The FO membrane achieved good pollutant removal efficiency during long-term operation. Distinct flux decline was also observed during the long-term filtration as a result of membrane fouling. The irreversible fouling gradually developed during the operation, and chemical cleaning was needed to more effectively recover membrane permeability.
Nomenclature A
overall water permeability
Ala
water permeability of fouling layer
Ame
water permeability of membrane
B
overall salt permeability
Bla
salt permeability of fouling layer
Bme
salt permeability of membrane
Ddraw
solute diffusion coefficient
Dml
diffusion coefficient of solutes within fouling layer
Jv
water flux
kCECP
mass transfer coefficient due to CECP
Km
mass transfer coefficient due to ICP
Km+foul
overall mass transfer coefficient due to both ICP and fouling mechanisms
Sme
membrane structure parameter
πdraw
osmotic pressure of the draw solution
πfeed
osmotic pressure of the feed solution
εla
porosity of fouling layer
tla
thickness of fouling layer 21
τla
tortuosity of fouling layer
εme
porosity of membrane support layer
tme
thickness of membrane support layer
τme
tortuosity of membrane support layer
Acknowledgments Financial support of this work by National Natural Science Foundation of China (51308400) and the National Science & Technology Pillar Program (2012BAJ21B05) is gratefully acknowledged. The instruments used in this research are supported by the Collaborative Innovation Center for Regional Environmental Quality. References [1] R.Z. Zhao, J.T. Novak, C.D. Goldsmith, Evaluation of on-site biological treatment for landfill leachates and its impact: A size distribution study, Water Res. 46 (2012) 3837-3848. [2] Z.F. Xie, Z.W. Wang, Q.Y. Wang, C.W. Zhu, Z.C. Wu, An anaerobic membrane bioreactor (AnDMBR) for landfill leachate treatment: Performance and microbial community identification, Bioresource Technol. 161 (2014) 29-39. [3] S. Kheradmand, A. Karimi-Jashni, M. Sartaj, Treatment of municipal landfill leachate using a combined anaerobic digester and activated sludge system, Waste Manage. 30 (2010) 1025-1031. [4]
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Hasar,
S.A.
Unsal, U.
Ipek,
S.
Karatas,
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Cinar,
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Yaman,
C.
Kinaci,
Stripping/flocculation/membrane bioreactor/reverse osmosis treatment of municipal landfill leachate, J. Hazard. Mater. 171 (2009) 309-317. [5] F.N. Ahmed, C.Q. Lan, Treatment of landfill leachate using membrane bioreactors: A review, Desalination 287 (2012) 41-54. [6] G. Insel, M. Dagdar, S. Dogruel, N. Dizge, E.U. Cokgor, B. Keskinler, Biodegradation characteristics and size fractionation of landfill leachate for integrated membrane treatment, J. Hazard. Mater. 260 (2013) 825-832. [7] R. Mahmoudkhani, A.H. Hassani, A. Torabian, S.M. Borghei, Study on High-strength Anaerobic Landfill Leachate Treatability By Membrane Bioreactor Coupled with Reverse Osmosis, Int. J. Environ. Res. 6 (2012) 129-138. [8] M. Campagna, M. Cakmakci, F.B. Yaman, B. Ozkaya, Molecular weight distribution of a full-scale landfill leachate treatment by membrane bioreactor and nanofiltration membrane, Waste Manage. 33 22
(2013) 866-870. [9] T.Y. Cath, A.E. Childress, M. Elimelech, Forward osmosis: Principles, applications, and recent developments, J. Membr. Sci. 281 (2006) 70-87. [10] S.F. Zhao, L. Zou, C.Y. Tang, D. Mulcahy, Recent developments in forward osmosis: Opportunities and challenges, J. Membr. Sci. 396 (2012) 1-21. [11] Y.H. Cho, J. Han, S. Han, M.D. Guiver, H.B. Park, Polyamide thin-film composite membranes based on carboxylated polysulfone microporous support membranes for forward osmosis, J. Membr. Sci. 445 (2013) 220-227. [12] Z.Z. Zhao, Study of membrane bioreactor plants for landfill leachate treatment, Low Carbon World 4 (2013) 94-96. [13] S. Loeb, L. Titelman, E. Korngold, J. Freiman, Effect of porous support fabric on osmosis through a Loeb-Sourirajan type asymmetric membrane, J. Membr. Sci. 129 (1997) 243-249. [14] J.S. Zhang, W.L.C. Loong, S.R. Chou, C.Y. Tang, R. Wang, A.G. Fane, Membrane biofouling and scaling in forward osmosis membrane bioreactor, J. Membr. Sci. 403 (2012) 8-14. [15] A. Tiraferri, N.Y. Yip, W.A. Phillip, J.D. Schiffman, M. Elimelech, Relating performance of thin-film composite forward osmosis membranes to support layer formation and structure, J. Membr. Sci. 367 (2011) 340-352. [16] E. Arkhangelsky, F. Wicaksana, S.R. Chou, A.A. Al-Rabiah, S.M. Al-Zahrani, R. Wang, Effects of scaling and cleaning on the performance of forward osmosis hollow fiber membranes, J. Membr. Sci. 415 (2012) 101-108. [17] J.R. McCutcheon, M. Elimelech, Modeling water flux in forward osmosis: Implications for improved membrane design, AICHE J. 53 (2007) 1736-1744. [18] A. Achilli, T.Y. Cath, A.E. Childress, Power generation with pressure retarded osmosis: An experimental and theoretical investigation, J. Membr. Sci. 343 (2009) 42-52. [19] W.C.L. Lay, T.H. Chong, C.Y. Tang, A.G. Fane, J.S. Zhang, Y. Liu, Fouling propensity of forward osmosis: investigation of the slower flux decline phenomenon, Water Sci. Technol. 61 (2010) 927-936. [20] W.C.L. Lay, J.S. Zhang, C.Y. Tang, R. Wang, Y. Liu, A.G. Fane, Factors affecting flux performance of forward osmosis systems, J. Membr. Sci. 394 (2012) 151-168. [21] Y.N. Wang, J. Wei, Q.H. She, F. Pacheco, C.Y. Tang, Microscopic characterization of FO/PRO membranes - A comparative study of CLSM, TEM and SEM, Environ. Sci. Technol. 46 (2012) 23
9995-10003. [22] M.Y. Chen, D.J. Lee, J.H. Tay, K.Y. Show, Staining of extracellular polymeric substances and cells in bioaggregates, Appl. Microbiol. Biotechnol. 75 (2007) 467-474. [23] Chinese NEPA, Water and Wastewater Monitoring Methods, 3rd ed., Chinese Environmental Science Publishing House, Beijing, China, 1997. [24] J.R. McCutcheon, R.L. McGinnis, M. Elimelech, Desalination by ammonia-carbon dioxide forward osmosis: Influence of draw and feed solution concentrations on process performance, J. Membr. Sci. 278 (2006) 114-123. [25] G.T. Gray, J.R. McCutcheon, M. Elimelech, Internal concentration polarization in forward osmosis: role of membrane orientation, Desalination 197 (2006) 1-8. [26] C.Y.Y. Tang, Q.H. She, W.C.L. Lay, R. Wang, A.G. Fane, Coupled effects of internal concentration polarization and fouling on flux behavior of forward osmosis membranes during humic acid filtration, J. Membr. Sci. 354 (2010) 123-133. [27] R.V. Linares, Z.Y. Li, M. Abu-Ghdaib, C.H. Wei, G. Amy, J.S. Vrouwenvelder, Water harvesting from municipal wastewater via osmotic gradient: An evaluation of process performance, J. Membr. Sci. 447 (2013) 50-56. [28] Y.L. Liu, B.X. Mi, Effects of organic macromolecular conditioning on gypsum scaling of forward osmosis membranes, J. Membr. Sci. 450 (2014) 153-161. [29] T. Maruyama, S. Katoh, M. Nakajima, H. Nabetani, T.P. Abbott, A. Shono, K. Satoh, FT-IR analysis of BSA fouled on ultrafiltration and microfiltration membranes, J. Membr. Sci. 192 (2001) 201-207. [30] E. Gomez-Ordonez, P. Ruperez, FTIR-ATR spectroscopy as a tool for polysaccharide identification in edible brown and red seaweeds, Food Hydrocolloids 25 (2011) 1514-1520. [31] H. Zhang, J.H. Qu, H.J. Liu, X. Zhao, Characterization of isolated fractions of dissolved organic matter from sewage treatment plant and the related disinfection by-products formation potential, J. Hazard. Mater. 164 (2009) 1433-1438. [32] H.Z. Ma, H.E. Allen, Y.J. Yin, Characterization of isolated fractions of dissolved organic matter from natural waters and a wastewater effluent, Water Res. 35 (2001) 985-996. [33] K.O. Agenson, T. Urase, Change in membrane performance due to organic fouling in nanofiltration (NF)/reverse osmosis (RO) applications, Sep. Purif. Technol. 55 (2007) 147-156. 24
[34] Y.L. Liu, B.X. Mi, Combined fouling of forward osmosis membranes: Synergistic foulant interaction and direct observation of fouling layer formation, J. Membr. Sci. 407 (2012) 136-144.
25
Table 1 Characteristics of the feed solution (n= 3). Items
Concentration (mg/L except for osmotic pressure)
COD
696±20
TOC
215±10
TDS
7100±47
TN
143±12
TP
0.3±0.0
Ca2+
925±22
Mg2+
324±14
Cl-
3676±58
Osmotic pressure (mOsm/kg)
785±4
26
Table 2 Virgin membrane characterization. Parameters
Value
Thickness of AL (μm)
9± 2
Thickness of SL (μm)
39± 6
Mean pore size of SL (μm)
5.2±0.9
Contact angle of AL (°)
90.6 ± 2.1
Contact angle of SL (°)
70.4 ± 0.7
Water permeability (A) (L/(m2 h bar))
1.33±0.12
Salt permeability (B) (L/(m2 h))
1.46±0.15
Salt rejection rate (%)
92.3±0.3
Note: AL: active layer; SL: support layer. The values are given as average ± standard deviation. The mean pore size of SL was obtained from the SEM image. Number of measurements (n): n=20 for thickness; n=100 for pore size; n=6 for contact angle; n=3 for the rest of the parameters.
27
Table 3 Water flux (Jv), mass transfer coefficient (Km) and solute diffusion coefficient (D) of the membrane using DI water as the feed solution. NaCl
Osmotic
concentration
pressure
Jv
Km
D
Jv
Km
D
(M)
(πdraw, bar)
(L/(m2 h))
(L/(m2 h))
(10-10m2/s)
(L/(m2 h)
(L/(m2 h)
(10-10m2/s)
0.5
25
9.8±0.2
8.71±0.32
9.51±0.34
14.2±0.2
5.37±0.10
5.87±0.10
1.0
50
13.7±0.1
9.13±0.11
9.97±0.12
26.0±0.3
7.74±0.11
8.45±0.12
1.5
75
19.5±0.1
12.38±0.12
13.52±0.13
30.2±0.2
7.77±0.06
8.49±0.06
2.0
100
21.9±0.2
12.51±0.13
13.66±0.14
39.2±0.2
9.38±0.04
10.24±0.04
2.5
125
24.4±0.1
13.05±0.08
14.25±0.09
44.4±0.2
10.01±0.05
10.93±0.05
3.0
150
26.9±0.2
13.74±0.12
15.00±0.13
48.4±0.3
10.41±0.07
11.36±0.08
4.0
200
28.0±0.2
12.69±0.13
13.86±0.14
51.8±0.3
10.37±0.06
11.32±0.07
AL-FS
AL-DS
Note: n=3 for Jv, Km and D; πfeed=0 bar, A=1.33 L/ (m2 h bar), B=1.46 L/ (m2 h), S=3.93×10-4 m for calculating Km and D using Eqs. (1)~(3).
28
Table 4 EDX microanalysis of fouled FO membrane. Weight percentage (%)
O
F
Na
Mg
Al
Si
P
Cl
K
Ca
Fe
Active layer
53.21
4.59
6.84
3.59
0.21
2.01
0.09
8.08
2.16
18.83
0.39
Support layer
62.86
--
15.56
--
--
--
--
19.76
--
1.82
--
Note: -- not detectable.
29
Table 5 Membrane permeability of the fouled membrane and the cleaned membrane (hydraulic and chemical cleaning). A
B
Ala 2
Bla
Jv
kCECP
Flux recovery
(L/(m2 h))
(%)
18.0
∞
-
2.03
8.2
6.75
-
2.01
2.49
16.0
65.59
88.9
20.78
22.22
17.8
72.73
98.9
(L/(m h
(L/(m
(L/(m h
(L/(m
(L/(m2
bar))
h)
bar))
h))
h))
Virgin membrane
1.33
1.46
∞
∞
Fouled membrane
0.72
0.85
1.57
0.80
0.92
1.25
1.37
Membrane samples
2
2
2
Membrane after hydraulic cleaning a Membrane after chemical cleaning a
Τhe value is for hydraulic cleaning performed at the end of the fourth run. πdraw=150 bar, πfeed=20.1 bar,
Km=13.74 L/(m2 h) for calculating kCECP by Eq. (4).
30
► An FO system is used for post-treatment of MBR-treated landfill leachate (LFL). ► AL-DS mode for FO leads to rapid fouling within 1 h for LFL treatment. ► About 98.6% of COD, 96.6% of TP, and 76.9% of ammonium are rejected by the FO system. ► Chemical cleaning is needed to remove irreversible fouling in long-term filtration.
31
Solution-diffusion theory AL-DS DI DI water AL-FS DI
ICP
AL-FS LFL AL-DS LFL
Landfill leachate
ICP/CECP Internal/external fouling
AL-FS filtration
Chemical cleaning
Irreversible fouling
Graphical abstract
32
A forward osmosis membrane system for the post-treatment of MBR-treated landfill leachate Ying Dong, Zhiwei Wang, Chaowei Zhu, Qiaoying Wang, Jixu Tang, Zhichao Wu
www.elsevier.com/locate/memsci
PII: DOI: Reference:
S0376-7388(14)00639-5 http://dx.doi.org/10.1016/j.memsci.2014.08.023 MEMSCI13135
To appear in:
Journal of Membrane Science
Received date: 28 May 2014 Revised date: 12 July 2014 Accepted date: 11 August 2014 Cite this article as: Ying Dong, Zhiwei Wang, Chaowei Zhu, Qiaoying Wang, Jixu Tang, Zhichao Wu, A forward osmosis membrane system for the posttreatment of MBR-treated landfill leachate, Journal of Membrane Science, http: //dx.doi.org/10.1016/j.memsci.2014.08.023 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting galley proof before it is published in its final citable form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
A forward osmosis membrane system for the post-treatment of MBR-treated landfill leachate Ying Dong1, Zhiwei Wang1,*, Chaowei Zhu2, Qiaoying Wang1, Jixu Tang1, Zhichao Wu1 1
State Key Laboratory of Pollution Control and Resource Reuse, School of Environmental
Science and Engineering, Tongji University, Siping Road 1239, Shanghai 200092, PR China 2
Chinese Research Academy of Environmental Sciences, Beijing 100012, PR China
Revised manuscript submitted to Journal of Membrane Science for possible publication
*Corresponding
Author.
Tel./fax:
+86-21-65980400
[email protected] (Z. Wang) 1
(Z.
Wang);
address:
Abstract Filtration behaviours, membrane fouling and cleaning were investigated in a forward osmosis (FO) membrane system used to post-treat the effluent of a membrane bioreactor (MBR) fed with landfill leachate (LFL). In short-term tests, it was observed that the water flux with the membrane active layer facing the draw solution (AL-DS mode) was lower than that with the membrane active layer facing the feed solution (AL-FS mode) for LFL treatment. Mathematical models could well simulate the flux evolution of AL-FS mode while the flux of AL-DS deviated from the modeling curve, suggesting that fouling could be rapidly developed within 1 h filtration for AL-DS mode. During long-term filtration, about 98.6% of COD, 96.6% of TP, and 76.9% of ammonium were rejected by the FO system. A decrease of water flux was also observed with an increase in operation time. Confocal laser scanning microscopy and Fourier transform infrared spectroscopy confirmed the existence of polysaccharides and proteins in the fouling layer. Inorganic fouling was mainly caused by Ca, Na, Mg, K, Si, Fe and Al. It was also found that the effect of cake enhanced concentration polarization played an important role during long-term operation. About 88.9% of the permeate flux was recovered after hydraulic cleaning while it reached 98.9% of the initial flux after chemical cleaning, indicating that chemical cleaning was needed to eliminate irreversible fouling and to recover membrane permeability during long-term operation. Keywords: Forward osmosis; landfill leachate; membrane fouling; membrane cleaning; wastewater treatment
1. Introduction Landfilling is still the most widely used method for municipal solid waste (MSW) disposal throughout the world, though recycling, composting and incineration are currently being encouraged [1, 2]. One of major concerns in landfills is the generation of leachate by the waste within the system and/or the infiltration of groundwater/rainfall. Landfill leachate (LFL) usually has high concentrations of organic pollutants, toxic materials (xenobiotic organics), ammonia, heavy metals and inorganic salts, and its characteristics can be significantly changed due to the variations of MSW composition, age, climatic conditions, landfill design and operational practice 2
[3]. If it is not properly collected and treated, it can potentially contaminate nearby surface and ground water. Development of efficient LFL treatment methods plays an important role in solving the potential threats of MSW landfills. To date, a number of treatment protocols have been proposed, e.g., physico-chemical methods including adsorption, stripping, coagulation, oxidation and membrane separation, and biological processes [4, 5]. For achieving high treatment efficiencies, physico-chemical treatment is always combined with biological processes. Physico-chemical units are placed either as pretreatment to reduce the loading rate for biological processes or as post-treatment to reach a high quality discharge standard. With the rapid development of membrane technology, combination of membrane bioreactor (MBR) with nano-filtration (NF) or reverse osmosis (RO) has become increasingly popular in LFL treatment [4, 6-8]. However, efforts should be taken to mitigate membrane fouling and reduce the energy consumption of NF/RO processes. Forward osmosis (FO) is an osmotically-driven membrane process, in which the driving force for separation is the difference in chemical potential between a concentrated draw solution and a broad range of aqueous solutions [9]. A semi-permeable FO membrane can retain solutes but allow water to transfer through the membrane. Compared to NF/RO, FO has several advantages such as no need for hydraulic pressure, high rejection of pollutants, and low membrane fouling propensity. In the past decade, FO has been used for oil and gas wastewater, and industrial and municipal wastewater treatment [10]. There is also a growing interest in the application of FO to LFL treatment [11]. The use of FO to replace NF/RO might be promising in addressing the existing obstacles. However, detailed information on the performance of FO in LFL treatment is scarce. In this study, an FO membrane system was used to post-treat the effluent of an MBR fed with landfill leachate. Batch experiments and long-term filtration behaviours were monitored in the FO membrane system. Membrane fouling and cleaning were also investigated. The obtained results are expected to provide a sound understanding of FO membrane performance in treating LFL.
3
2. Materials and methods 2.1. Experimental set-up A lab-scale FO system, as shown in Fig. 1, was used in this study. The membrane cell comprises equally structured rectangular channels on both sides of the membrane. The dimensions of each channel are 100 mm long, 20 mm wide and 1.5 mm deep. The effective membrane area is approximately 0.002 m2. Two variable-speed diaphragm pumps (Mini-pump, Benquan, Shanghai, China) were used to recirculate the feed and draw solutions, respectively. A cross flow velocity of 25 cm/s was maintained on both sides of the membrane through recirculation. The water flux was determined by measuring the weight changes of the feed solution in real time with an electronic balance (Ohaus, New Jersey, USA) connected to a laptop and a data logging system. All experiments were carried out at room temperature (25 ± 0.5 oC). Before use, the membrane was soaked in DI water for 24 hours and further washed with DI water under a cross-flow velocity 25 cm/s for 30 minutes. The volumes of the feed and draw solution tanks are both 6 L. The FO membrane used in this study was provided by Hydration Technologies Innovation (HTI, Albany, USA). It is made of cellulose triacetate (CTA) with embedded polyester mesh as mechanical support.
Fig. 1. Schematic of the lab-scale FO system. 2.2. Operating conditions 2.2.1. Short-term filtration Since the membrane has an asymmetric structure, the water flux of both membrane orientations (i.e., AL-FS with the membrane active layer facing the feed solution, and AL-DS with the membrane active layer facing the draw solution) was determined. Firstly, DI water was used as 4
feed solution to obtain the DI water flux using different concentrations of draw solutions. Then, the effluent from an MBR fed with LFL was used as feed solution. The characteristics of the feed solution are shown in Table 1, and the detailed information on this MBR can be found elsewhere [12]. Each batch test lasted for 1 hour. After each test, the feed and draw solutions were changed to DI water, and the membrane was cleaned for 30 min under the cross-flow velocity of 25 cm/s. Analytical grade sodium chloride (NaCl) was used to prepare draw solutions with concentrations ranging from 0.5 M to 4.0 M.
Table 1 Characteristics of the feed solution (n= 3). Items
Concentration (mg/L except for osmotic pressure)
COD
696±20
TOC
215±10
TDS
7100±47
TN
143±12
TP
0.3±0.0
2+
925±22
Mg2+
324±14
Cl-
3676±58
Osmotic pressure (mOsm/kg)
785±4
Ca
2.2.2. Long-term filtration During the long-term filtration experiment, 3.0 M NaCl was used as the draw solution and the MBR-treated effluent was adopted as the feed solution. The feed and draw solutions were replaced with fresh ones every 24 hours in order to avoid significant changes in their concentrations. The AL-FS orientation was used in order to avoid dramatic flux decline. No cleaning was performed until the water flux was decreased to about 60% of the initial. Hydraulic cleaning was carried out for 1 h under the cross-flow velocity of 25 cm/s by changing the feed and draw solutions to DI water. After 5 runs, the membrane was taken out from the cell, and 1-h chemical cleaning under the cross flow velocity of 25 cm/s was conducted using 1% Alconox (White Plains, USA) solution as the feed and draw solution. Afterwards, the feed and draw solution was changed to DI water for 5
10 min under cross flow in order to remove the remaining Alconox solution. Another run was performed after the chemical cleaning to observe the flux recovery using the MBR-treated effluent as feed solution. A parallel experiment was carried out in order to obtain the fouled membrane samples for fouling characterization as mentioned in Section 2.4. 2.3. Modeling FO performance An analytical model, incorporating the effect of internal concentration polarization (ICP), has been developed by Loeb et al. [13] for evaluating FO performance. According to this model, the FO water flux (Jv) can be worked out by Eq. (1) for the AL-FS orientation and Eq. (2) for the AL-DS orientation [14]. ⎛ Aπ draw + B ⎞ J v = K m ln ⎜ ⎟ ⎝ Aπ feed + J v + B ⎠
(1)
⎛ Aπ − Jv + B ⎞ J v = K m ln ⎜ draw ⎟ A π feed + B ⎝ ⎠
(2)
In Eqs. (1) and (2), πdraw and πfeed are the osmotic pressure of the draw solution and feed solution, respectively. A and B are the water and salt permeability of the rejection layer, respectively, which can be determined by RO filtration tests as described by Tiraferri et al. [15]. Km, the mass transfer coefficient, is related to the ICP phenomenon within the porous support layer. Km can be worked out using the solute diffusion coefficient Ddraw divided by the membrane structure parameter Sme.
Km =
Ddraw Ddraw ⋅ ε me = S me tme ⋅ τ me
(3)
where εme, tme and τme are the porosity, thickness and tortuosity of the membrane support layer, respectively. The diffusion coefficient (D) can be worked out through Eq. (3) using Km obtained from Eq. (1) and Eq. (2) and the membrane structure parameter S=3.93×10-4 m [16]. Eqs. (1) and (2) are valid for well-defined feed and draw solutions in most cases [17, 18]; however, the effects of fouling under real operating conditions are not explicitly accounted for. To address this issue, Lay et al. [19] proposed a fouling-incorporated water flux model for an FO system based on a resistance-in-series approach for a fouling condition with cake enhanced concentration polarization (CECP). The water flux of an FO system with the AL-FS orientation 6
can be expressed by Eq. (4).
B⎞ ⎡⎛ J v = A ⋅ ⎢⎜ π draw + ⎟ ⋅ e −( J v A⎠ ⎝ ⎣
Km )
B⎞ ⎛ − ⎜ π feed + ⎟ ⋅ e( J v A⎠ ⎝
kCECP )
⎤ (4) ⎥ ⎦
where A and B are the overall water and salt permeability coefficients, which are related to the coefficients of a membrane (subscript ‘me’) and fouling layer (subscript ‘la’) as follows [19].
1 1 1 = + A Ame Ala
(5)
1 1 1 = + B Bme Bla
(6)
The coefficient kCECP can be expressed via a mass transfer coefficient in Eq. (7) [19].
kCECP =
Dml ⋅ ε la δ la ⋅ τ la
(7)
where Dml is the diffusion coefficient on average of the solutes within the fouling layer, and εla, δla, and τla are the porosity, thickness and tortuosity of the fouling layer, respectively. A smaller kCECP indicates that the CECP effects are more significant. If kCECP=∞, the CECP effects are negligible. The above-mentioned Eq. (4) can be rewritten as Eq. (1) by setting kCECP=∞. For AL-DS orientation in an FO system, the water flux can be modeled using Eq. (8) under the assumption that the external concentration effects are negligible.
B⎞ ⎛ B⎞ ⎡⎛ J v = A ⋅ ⎢⎜ π draw + ⎟ − ⎜ π feed + ⎟ ⋅ e( J v A A⎠ ⎝ ⎠ ⎝ ⎣
K m+foul )
⎤ (8) ⎥ ⎦
where the parameter Km+foul is a complex overall mass transfer coefficient that couples both ICP and fouling mechanisms. In the absence of fouling or other performance diminishing effects, Km+foul is equal to Km [20], and Eq. (8) can be reduced back to Eq. (2). In this study, Eqs. (1) and (2) were used to model the water flux under AL-FS and AL-DS orientations respectively in the batch experiments. Eq. (4) and Eq. (8) were also employed to evaluate the development of fouling in the bath filtration for LFL treatment. For the long-term experiments under AL-FS orientation, Eq. (4) was adopted to assess the filtration behaviour.
7
2.4. Analytical methods 2.4.1. Morphological analyses For scanning electron microscopy (SEM) analysis, a scanning electron microscope (Model XL 30, Philips, Netherlands) equipped with an energy-dispersive X-ray spectroscope (EDX) was used. Before observation, all samples were dried in vacuum at room temperature (25 °C), and sputter-coated with a thin layer of gold. The fouled FO membrane was also observed. Elemental analysis of the foulant on the membrane surface was carried out using EDX immediately after SEM observation. A confocal laser scanning microscope (CLSM) (Nikon A1, Tokyo, Japan) was used in both fluorescent mode and reflective mode to visualize the membranes. The samples can be stained using the methods described by Wang et al. [21] and Chen et al. [22]. Lateral x-y scans were performed with a ×20 objective, and images were stored with a resolution of 512 × 512 pixels (representing an area of 360 × 360 μm). The scans were conducted from the active layer (active layer facing the objective lens). Images were collected at a series of sampling depths with a scanning depth increment of 1.65 μm in the z-direction to obtain the internal structure of membrane samples. For this study, we set z = 0 at the membrane active layer surface, with the positive z-direction pointing toward the supporting layer [21]. SEM and CLSM images were analyzed using image analysis software (ImageJ, NIH, USA) to obtain the information of pore size and porosity of the virgin membrane samples. The CLSM and SEM images were converted into 8-bit gray scale images, which were then transformed to black and white images by adjusting the threshold value. The porosity was determined as the ratio of the total void area to the gross area analyzed using the software. The sampled area for each analysis was about 100 × 100 μm for CLSM images, and about 70 × 88 μm for SEM images. In both CLSM and SEM, at least 3 images of each membrane were determined [21] . 2.4.2. Fourier transform infrared spectroscopy (FTIR) For FTIR studies, fouled membrane collected from the long-term experiment was pretreated in a vacuum freeze drying machine for 24 hours. Then the dried membrane together with the foulant was directly analyzed by FTIR spectroscopy with an OMNI sampler and an attenuated total 8
reflectance (ATR) accessory (Nicolet 5700, Thermo Electron Corporation, USA), in order to determine the functional groups of the foulants. The FTIR of the clean membrane was used as the blank, and its transmittance was subtracted from the spectra of the fouled membrane. The FTIR spectra were analyzed based on the Sadtler handbook of Infrared Spectra. 2.4.3. Wastewater characteristics Analyses of chemical oxygen demand (COD), total nitrogen (TN), and total phosphate (TP) were performed based on the standard Chinese NEPA methods [23]. The concentrations of Ca and Mg
were
determined
using
an
inductively
coupled
plasma
emission
spectrometer
(ICP-Agilent720ES, Agilent Technologies, Santa Clara, USA). Total dissolved solids (TDS) were measured using a DDSJ-308A conductivity meter (Leici, Shanghai, China). Total organic carbon (TOC) was determined by using a VCPN total organic carbon analyzer (Shimadzu Corporation, Kyoto, Japan). The measurement of osmotic pressure was carried out by using an FM-9J osmometer (Yida, Shanghai, China). The rejection of pollutants in the FO system is obtained from Eq. (9). R=
VC ×100% V0C0
(9)
In Eq. (9), V0 and V are the initial volume of the feed solution and the final volume of the feed solution prior to the change into fresh one every day, respectively. C0 and C are the initial and final concentrations of the pollutants measured in the feed solution, respectively. The rejection values of pollutants are reported as the average values and standard deviations during 36-d operation.
3. Results and discussion 3.1. Osmosis membrane characterization and short-term filtration The characteristics of the virgin FO membrane are summarized in Table 2. The FO membrane, as shown in Fig. S1 (see the supporting information), has a dense active layer (9 μm) and a support layer (39 μm). The polyester (PE) mesh fibers at a spacing of about 110 μm could be clearly seen from the cross section and the support layer of the membrane. A larger fiber-to-fiber distance of about 160 μm has been also reported for other membranes [21]. The total thickness of 9
the FO membrane is 48 μm, much thinner than standard RO membranes in order to enhance FO membrane permeability [9]. The membrane’s support layer (Fig. S1) has a mean pore size of 5.2 μm. The relatively larger pore size can mitigate internal concentration polarization phenomenon. It has been also reported that the improvement of hydrophilicity can facilitate the FO membrane permeability [9]. However, the AL of the virgin membrane used in this study is hydrophobic, indicating there is much space to further modify the membrane. The water permeability (A) and salt permeability (B) of the membrane are consistent with the results of previous publications [20, 24, 25]. Table 2 Virgin membrane characterization. Parameters
Value
Thickness of AL (μm)
9± 2
Thickness of SL (μm)
39± 6
Mean pore size of SL (μm)
5.2±0.9
Contact angle of AL (°)
90.6 ± 2.1
Contact angle of SL (°)
70.4 ± 0.7
Water permeability (A) (L/(m2 h bar))
1.33±0.12
Salt permeability (B) (L/(m2 h))
1.46±0.15
Salt rejection rate (%)
92.3±0.3
Note: AL: active layer; SL: support layer. The values are given as average ± standard deviation. The mean pore size of SL was obtained from the SEM image. Number of measurements (n): n=20 for thickness; n=100 for pore size; n=6 for contact angle; n=3 for the rest of the parameters.
In order to further examine the inner structure of the membrane, layer-by-layer CLSM scanning was carried out. The CLSM images and corresponding pore size and porosity of the inner structure are illustrated in Fig. 2. There are no visible pores in the active layer (z=0 μm), which might be due to the dense property of the active layer and the low magnification of CLSM. Smaller pores (z<10 μm) were found near the active layer, and larger pores (z>10 μm) were observed on the support layer side, which is consistent with the results reported by Wang et al. [21]. It is interesting to find that the inner structure of the SL has a larger pore size and porosity than the surface of the SL (z=49 μm). This may increase the difficulty to eliminate ICP through increasing cross-flow velocity on the SL side. 10
(a) (c) (b) 100 μm 100 μm 100 μm (f) (d) (e) 16
100 μm
100 μm
Pore diameter (μm)
14
Pore diameter Porosity
12 10
100 90 80 70 60 50 40 30 20 10 0
Porosity (%)
8 6 4 2 0 0 10 20 30 40 50 Distance from the AL surface (μm)
Fig. 2. CLSM images of a clean FO membrane viewed from the active layer side. (a) z=0 μm; (b) z=8.25 μm; (c) z=16.5 μm; (d) z=24.75 μm; (e) z=33 μm; and (f) the pore size and porosity of the membrane inner structure as a function of z (the distance from the membrane active layer surface).
Table 3 summarizes the water flux, mass transfer coefficient and solute diffusion coefficient of the virgin membrane using DI water as the feed solution. A strong correlation between the water flux and draw solution concentration was observed, which was due to the increasing osmotic pressure difference across the membrane. However, the flux was not evidently changed at higher NaCl solution concentration, suggesting that the water flux is non-linearly correlated with the draw solution concentration as a result of ICP [17]. The effect of ICP is supposed to be much severer at higher water fluxes [20, 24]. The membrane orientation has a significant effect on the water flux. The water flux of DI water in AL-DS mode is much higher than that in AL-FS mode with the same draw solution concentration.
11
Table 3 Water flux (Jv), mass transfer coefficient (Km) and solute diffusion coefficient (D) of the membrane using DI water as the feed solution. NaCl
Osmotic
concentration
pressure
(M)
(πdraw, bar)
(L/(m h))
(L/(m h))
0.5
25
9.8±0.2
8.71±0.32
1.0
50
13.7±0.1
1.5
75
2.0
AL-FS Jv 2
AL-DS
Km 2
D -10
Jv 2
(10 m /s)
Km 2
2
D -10
(10 m2/s)
(L/(m h)
(L/(m h)
9.51±0.34
14.2±0.2
5.37±0.10
5.87±0.10
9.13±0.11
9.97±0.12
26.0±0.3
7.74±0.11
8.45±0.12
19.5±0.1
12.38±0.12
13.52±0.13
30.2±0.2
7.77±0.06
8.49±0.06
100
21.9±0.2
12.51±0.13
13.66±0.14
39.2±0.2
9.38±0.04
10.24±0.04
2.5
125
24.4±0.1
13.05±0.08
14.25±0.09
44.4±0.2
10.01±0.05
10.93±0.05
3.0
150
26.9±0.2
13.74±0.12
15.00±0.13
48.4±0.3
10.41±0.07
11.36±0.08
4.0
200
28.0±0.2
12.69±0.13
13.86±0.14
51.8±0.3
10.37±0.06
11.32±0.07
Note: n=3 for Jv, Km and D; πfeed=0 bar, A=1.33 L/ (m2 h bar), B=1.46 L/ (m2 h), S=3.93×10-4 m for calculating Km and D using Eqs. (1)~(3).
In order to further compare the water flux of FO membranes using DI water and LFL as the feed solutions, Fig. 3 is drawn to illustrate the obtained experimental flux and the modeled flux as functions of osmotic pressure difference. It is evident that the water flux of FO membranes deviates from linear curve (Jv=A(πdraw-πfeed)) based on the classical solution-diffusion theory [9], indicating that the fouling (e.g., ICP using DI as feed solution and complex fouling due to the use of LFL as feed solution) can significantly impact the water fluxes in real applications [13]. It was also found that the mathematic models, as expressed by Eqs. (1) and (2), could well fit the experimental data in both AL-FS and AL-DS modes using DI water as feed solution. The flux deviation from the linear correlation can be well explained by dilutive ICP in the SL of AL-FS mode and concentrative ICP in AL-DS mode. It has been reported that A has greater impacts on water flux than Km with smaller osmotic driving force (e.g., (πdraw-πfeed) ≤10 bar); however, Km dominates the influence with greater driving force ((πdraw-πfeed) ≥20 bar) [20]. The membrane used in this study has a higher Km value and a similar A compared to the membranes in the reference [20], suggesting that higher membrane fluxes could be achieved in particular in the range of larger driving force. 12
For LFL, it is observed that the water flux in the AL-FS mode is higher than that in the AL-DS mode during the short-term filtration (1 h). Interestingly, this is contrary to the results of using activated sludge as the feed solution [26]. To elucidate the mechanisms, we used Eqs. (7) and (8) to simulate the water fluxes of LFL using AL-FS and AL-DS modes, respectively. It can be seen that the experimental data for AL-FS mode is well modeled by Eq. (7) by setting kCECP= ∞ and
πfeed= 20.1 bar (see Fig. 3). An infinite (∞) value of kCECP indicates that the CECP effect is negligible [20]. The results suggested that the CECP was not obviously formed in the AL-FS mode for LFL treatment in short-term filtration. However, it was interestingly found that the experimental fluxes were much lower than the modeled values under the AL-DS mode for LFL treatment by using Eq. (8) (setting Km+foul= Km and πfeed= 20.1 bar), demonstrating that the parameter Km+foul could not be simplified as Km. The parameter Km+foul is considered as a complex overall mass transfer coefficient that incorporates both ICP and fouling processes in the AL-DS mode [20]. Take the DS concentration of 3.0 M as an example, the Km+foul for LFL calculated from Eq. (8) is 5.33 L/(m2 h), while the value of Km for DI water is 10.41 L/ (m2 h). This indicates that the fouling of FO membranes in the AL-DS mode can be rapidly developed within 1 h filtration. Therefore, for real wastewater treatment, the AL-FS mode is always adopted in long-term filtration. Compared with the water flux of using activated sludge as the feed solution under the same experimental conditions (DS: 0.5 M NaCl) [26], the water flux of LFL was much lower. It demonstrates that the fouling conditions of LFL are more critical, which can be attributed to the more complicated composition and the higher salts in LFL.
13
AL-DS DI
AL-FS DI
AL-FS LFL AL-DS LFL
Fig. 3. Modeled and experimental water flux as a function of osmotic pressure difference. symbols: □ experimental data of water flux using AL-DS mode and feed solution DI water; ■ experimental data of water flux using AL-FS mode and feed solution DI water;● experimental data of water flux using AL-FS mode and feed solution (LFL); ○ experimental data of water flux using AL-DS mode and feed solution LFL. In calculating the modeled water flux, πfeed=0 bar for DI water, πfeed=20.1 bar for landfill leachate (LFL), A=1.33 L/ (m2 h bar), and B= 1.46 L/ (m2 h); the values of Km were obtained by fitting the measured data (Km=2.513ln(Δπ)+0.5475, R2=0.8076 in AL-FS mode; Km=2.5516ln(Δπ)−2.6481, R2=0.9481 in AL-DS mode). The red solid line is the modeled flux using Jv=A(πdraw-πfeed), and other lines indicate the modeled flux using Eqs. (1) and (2) based on the adopted membrane orientation.
3.2. Membrane separation performance during long-term operation Fig. 4 shows the water flux evolution of LFL in AL-FS mode during 36-day operation using the AL-FS mode. The initial flux at the beginning of the experiment was 18.0 L/ (m2 h), which is in agreement with the experimental water flux as shown in Fig. 3 during the short-term filtration. Since both the feed and draw solutions were replaced with fresh ones every day in this study, the impact of concentration changes on water flux was limited. However, distinct flux decline during each run can be observed, indicating that fouling was a critical issue in the long term FO 14
performance [14]. Hydraulic cleaning was carried out in the first 4 runs (indicated by the blue circles), while chemical cleaning (red circle) was performed at the end of fifth run. It is also evident in Fig. 4 that irreversible fouling is gradually developed during the long-term filtration as revealed by the dashed line, indicating that chemical cleaning should be periodically carried out in order to sustain the operation.
Fig. 4. Water flux during the long-term filtration experiment under the AL-FS mode for LFL treatment.
Fig. 5 shows the average retention efficiency of the pollutants in the LFL. About 98.6% of COD, 96.6% of TP, and 76.9% of ammonium were rejected by the FO system. The rejection of Ca2+ and Mg2+ was also very effective. The lower rejection of ammonium compared with Ca2+ and Mg2+ should be attributed to its smaller ionic radius or hydrated ionic radius. In addition, Ca2+ and Mg2+ could also result in the scaling on the membrane surface [16, 27].
15
Fig. 5. The rejection of pollutants during the long-term filtration experiment (n=36). 3.3. Membrane fouling and cleaning In order to examine the fouling behaviours, the fouled membrane in the long-term filtration experiment was taken out from the membrane cell after the fifth fouling test run without any cleaning and analyzed by SEM and CLSM. Fig. 6 (a) and Fig. 6 (b) show the SEM images of the active and support layer of the fouled membrane, respectively. As shown in Fig. 6 (a), an obvious deposition of foulants on the active layer was observed, while the fouling was not severe on the support layer (Fig. 6 (b)). Unlike a cake layer formed on membrane surfaces in an FO-MBR [14], a number of crystallization could be found on the active layer, indicating that the dominant foulants are different in various FO systems. The EDX results (Table 4) show that Ca, Na, Mg, K, Si, Fe and Al are the main inorganic elements in the foulants, especially calcium, which is considered to have negative effects on the membrane performance [28]. (a)
(b)
200 μm
200 μm
Fig. 6. SEM images of fouled FO membrane surfaces (a) SEM image of active layer; (b) SEM image of support layer.
16
The components of the foulants on the active and support layers were also analyzed using FTIR. Fig. 7 shows the FTIR spectra of the fouling layer on both sides. The broad peak at 1045 cm−1, 1232 cm−1 and 1742 cm−1 are associated with polysaccharides or polysaccharide-like substances. The peak at 1233 cm−1 is assigned to the presence of sulfate ester groups which is a characteristic component in fucoidan and sulphated polysaccharides, while the peak at 1742 cm−1 is assigned to the carboxylic acid ester form (C=O) in alginic acid standard [29]. The broad peak at 1045 cm−1 is also related to polysaccharides or polysaccharide-like substances [30]. Signals at 1515 cm−1 correspond to the amide II bands [29], indicating the presence of proteins. The FTIR results indicate that protein and polysaccharide substances may be the major organic foulants.
Table 4 EDX microanalysis of fouled FO membrane. Weight percentage (%)
O
F
Na
Mg
Al
Si
P
Cl
K
Ca
Fe
Active layer
53.21
4.59
6.84
3.59
0.21
2.01
0.09
8.08
2.16
18.83
0.39
Support layer
62.86
--
15.56
--
--
--
--
19.76
--
1.82
--
Note: -- not detectable.
Fig. 7. FTIR spectra of the membrane foulants in the long-term fouling experiment
For further examining the fouling layer on the active layer surface of the FO membrane, fluorescent staining and CLSM observation were performed. Figs. 8 (a)-(c) demonstrate the distributions of α-polysaccharides (Con A), β-polysaccharides (Calcofluor white) and proteins (FITC). According to the fluorescent intensity presented in Fig. 8, abundant proteins and polysaccharides (especially α-polysaccharides) exist in the fouling layer. They have been reported 17
to be the major components of organic fouling during membrane filtration of water and wastewater [31-33]. In combination with EDX analysis, the fouling could be interpreted as the comprehensive impacts of organic foulants, inorganic precipitates and/or organic-inorganic complexes. The formation of inorganic scaling might be ascribed to the severe effects of concentration polarization caused by inorganic ions existing in the LFL and the high rejection property of the selective layer of the FO membrane. Furthermore, the organic fouling layer formed on the membrane surface may have hindered the diffusion of inorganic foulants back to the feed solution, promoting the inorganic precipitation. It is suggested that synergistic effects between organic and inorganic foulants can result in much severer fouling of FO membranes [34]. (a)
(b)
100 μm
100 μm
(c)
(d)
100 μm
100 μm
Fig. 8. The CLSM images of stained fouling layer (a) CLSM image of α-polysaccharides (Con A); (b) CLSM image of β-polysaccharides (Calcofluor white); (c) CLSM image of proteins (FITC); (d) combination of the individual images.
Table 5 summarizes the membrane permeability of fouled and cleaned membrane during the long-term filtration. A dramatic decline of A and B of the fouled membrane was observed compared with the virgin membrane. Hydraulic cleaning led to an increase of A and B, and chemical cleaning further recovered the value almost back to the initial. This suggests that the effects of organic foulants could not be ignored in this study except for inorganic fouling [14]. Ala 18
and Bla can be worked out through Eqs. (5) and (6), respectively, and their values in Table 5 clearly show that the decrease of A and B is associated with the decrease of Ala and Bla from the infinite to a certain value due to fouling. Higher values of Ala and Bla can result in larger A and B. The parameter kCECP can indicate the fouling condition; a lower value denotes a severer fouling. It clearly shows the CECP plays an important role in the fouling process during the long-term operation. Regarding the water flux recovery, about 88.9% of the permeate flux was recovered after the hydraulic cleaning while the value reached 98.9% of the initial after the chemical cleaning. This indicates that chemical cleaning is needed to address irreversible membrane fouling and to recover membrane permeability during long-term operation.
Table 5 Membrane permeability of the fouled membrane and the cleaned membrane (hydraulic and chemical cleaning). B
A
Ala 2
Bla 2
Jv
(L/(m h
(L/(m
(L/(m h
(L/(m
(L/(m2
bar))
h)
bar))
h))
h))
Virgin membrane
1.33
1.46
∞
∞
Fouled membrane
0.72
0.85
1.57
0.80
0.92
1.25
1.37
Membrane samples
2
2
kCECP 2
Flux recovery
(L/(m h))
(%)
18.0
∞
-
2.03
8.2
6.75
-
2.01
2.49
16.0
65.59
88.9
20.78
22.22
17.8
72.73
98.9
Membrane after hydraulic cleaning a Membrane after chemical cleaning a
Τhe value is for hydraulic cleaning performed at the end of the fourth run. πdraw=150 bar, πfeed=20.1 bar,
Km=13.74 L/(m2 h) for calculating kCECP by Eq. (4).
It has to be pointed out that daily replacement of both draw and feed solutions was performed in order to avoid the decline of the driving force (πdraw –πfeed) in this study. Under this circumstance, the water recovery rate is very low. For practical applications, the water recovery rate needs reaching a high value, and also continuous flow mode should be adopted. For instance, the continuous inflow of the influent could ensure that the concentration of the feed solution is maintained at relatively constant value. The draw solution can be also kept constant through adding more concentrated salt solutions or concentrating draw solution using other devices (e.g., 19
reverse osmosis). Therefore, the obtained results in this study can reflect some practical cases that the concentration of the feed solution is maintained at about 700 mg/L in terms of COD. Further studies are needed in order to simulate more practical conditions of using FO to treat landfill leachate.
20
4. Conclusions An FO membrane system was used for post-treatment of landfill leachate from a membrane bioreactor, and fouling behaviours, membrane fouling and cleaning were studied. In short-term batch experiments, the water flux in the AL-FS mode was higher than that in the AL-DS mode, indicating that the fouling of FO membranes in the AL-DS mode was rapidly developed within 1 h filtration. The FO membrane achieved good pollutant removal efficiency during long-term operation. Distinct flux decline was also observed during the long-term filtration as a result of membrane fouling. The irreversible fouling gradually developed during the operation, and chemical cleaning was needed to more effectively recover membrane permeability.
Nomenclature A
overall water permeability
Ala
water permeability of fouling layer
Ame
water permeability of membrane
B
overall salt permeability
Bla
salt permeability of fouling layer
Bme
salt permeability of membrane
Ddraw
solute diffusion coefficient
Dml
diffusion coefficient of solutes within fouling layer
Jv
water flux
kCECP
mass transfer coefficient due to CECP
Km
mass transfer coefficient due to ICP
Km+foul
overall mass transfer coefficient due to both ICP and fouling mechanisms
Sme
membrane structure parameter
πdraw
osmotic pressure of the draw solution
πfeed
osmotic pressure of the feed solution
εla
porosity of fouling layer
tla
thickness of fouling layer 21
τla
tortuosity of fouling layer
εme
porosity of membrane support layer
tme
thickness of membrane support layer
τme
tortuosity of membrane support layer
Acknowledgments Financial support of this work by National Natural Science Foundation of China (51308400) and the National Science & Technology Pillar Program (2012BAJ21B05) is gratefully acknowledged. The instruments used in this research are supported by the Collaborative Innovation Center for Regional Environmental Quality. References [1] R.Z. Zhao, J.T. Novak, C.D. Goldsmith, Evaluation of on-site biological treatment for landfill leachates and its impact: A size distribution study, Water Res. 46 (2012) 3837-3848. [2] Z.F. Xie, Z.W. Wang, Q.Y. Wang, C.W. Zhu, Z.C. Wu, An anaerobic membrane bioreactor (AnDMBR) for landfill leachate treatment: Performance and microbial community identification, Bioresource Technol. 161 (2014) 29-39. [3] S. Kheradmand, A. Karimi-Jashni, M. Sartaj, Treatment of municipal landfill leachate using a combined anaerobic digester and activated sludge system, Waste Manage. 30 (2010) 1025-1031. [4]
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Table 1 Characteristics of the feed solution (n= 3). Items
Concentration (mg/L except for osmotic pressure)
COD
696±20
TOC
215±10
TDS
7100±47
TN
143±12
TP
0.3±0.0
Ca2+
925±22
Mg2+
324±14
Cl-
3676±58
Osmotic pressure (mOsm/kg)
785±4
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Table 2 Virgin membrane characterization. Parameters
Value
Thickness of AL (μm)
9± 2
Thickness of SL (μm)
39± 6
Mean pore size of SL (μm)
5.2±0.9
Contact angle of AL (°)
90.6 ± 2.1
Contact angle of SL (°)
70.4 ± 0.7
Water permeability (A) (L/(m2 h bar))
1.33±0.12
Salt permeability (B) (L/(m2 h))
1.46±0.15
Salt rejection rate (%)
92.3±0.3
Note: AL: active layer; SL: support layer. The values are given as average ± standard deviation. The mean pore size of SL was obtained from the SEM image. Number of measurements (n): n=20 for thickness; n=100 for pore size; n=6 for contact angle; n=3 for the rest of the parameters.
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Table 3 Water flux (Jv), mass transfer coefficient (Km) and solute diffusion coefficient (D) of the membrane using DI water as the feed solution. NaCl
Osmotic
concentration
pressure
Jv
Km
D
Jv
Km
D
(M)
(πdraw, bar)
(L/(m2 h))
(L/(m2 h))
(10-10m2/s)
(L/(m2 h)
(L/(m2 h)
(10-10m2/s)
0.5
25
9.8±0.2
8.71±0.32
9.51±0.34
14.2±0.2
5.37±0.10
5.87±0.10
1.0
50
13.7±0.1
9.13±0.11
9.97±0.12
26.0±0.3
7.74±0.11
8.45±0.12
1.5
75
19.5±0.1
12.38±0.12
13.52±0.13
30.2±0.2
7.77±0.06
8.49±0.06
2.0
100
21.9±0.2
12.51±0.13
13.66±0.14
39.2±0.2
9.38±0.04
10.24±0.04
2.5
125
24.4±0.1
13.05±0.08
14.25±0.09
44.4±0.2
10.01±0.05
10.93±0.05
3.0
150
26.9±0.2
13.74±0.12
15.00±0.13
48.4±0.3
10.41±0.07
11.36±0.08
4.0
200
28.0±0.2
12.69±0.13
13.86±0.14
51.8±0.3
10.37±0.06
11.32±0.07
AL-FS
AL-DS
Note: n=3 for Jv, Km and D; πfeed=0 bar, A=1.33 L/ (m2 h bar), B=1.46 L/ (m2 h), S=3.93×10-4 m for calculating Km and D using Eqs. (1)~(3).
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Table 4 EDX microanalysis of fouled FO membrane. Weight percentage (%)
O
F
Na
Mg
Al
Si
P
Cl
K
Ca
Fe
Active layer
53.21
4.59
6.84
3.59
0.21
2.01
0.09
8.08
2.16
18.83
0.39
Support layer
62.86
--
15.56
--
--
--
--
19.76
--
1.82
--
Note: -- not detectable.
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Table 5 Membrane permeability of the fouled membrane and the cleaned membrane (hydraulic and chemical cleaning). A
B
Ala 2
Bla
Jv
kCECP
Flux recovery
(L/(m2 h))
(%)
18.0
∞
-
2.03
8.2
6.75
-
2.01
2.49
16.0
65.59
88.9
20.78
22.22
17.8
72.73
98.9
(L/(m h
(L/(m
(L/(m h
(L/(m
(L/(m2
bar))
h)
bar))
h))
h))
Virgin membrane
1.33
1.46
∞
∞
Fouled membrane
0.72
0.85
1.57
0.80
0.92
1.25
1.37
Membrane samples
2
2
2
Membrane after hydraulic cleaning a Membrane after chemical cleaning a
Τhe value is for hydraulic cleaning performed at the end of the fourth run. πdraw=150 bar, πfeed=20.1 bar,
Km=13.74 L/(m2 h) for calculating kCECP by Eq. (4).
30
► An FO system is used for post-treatment of MBR-treated landfill leachate (LFL). ► AL-DS mode for FO leads to rapid fouling within 1 h for LFL treatment. ► About 98.6% of COD, 96.6% of TP, and 76.9% of ammonium are rejected by the FO system. ► Chemical cleaning is needed to remove irreversible fouling in long-term filtration.
31
Solution-diffusion theory AL-DS DI DI water AL-FS DI
ICP
AL-FS LFL AL-DS LFL
Landfill leachate
ICP/CECP Internal/external fouling
AL-FS filtration
Chemical cleaning
Irreversible fouling
Graphical abstract
32