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Phosphate removal from domestic wastewater using thermally modified steel slag Jian Yu1,3 , Wenyan Liang1,⁎, Li Wang1 , Feizhen Li1 , Yuanlong Zou2 , Haidong Wang2 1. College of Environmental Science and Engineering, Beijing Forestry University, Beijing 100083, China. E-mail: [email protected] 2. Central Research Institute of Building and Construction Co., Ltd., MCC, Beijing 100088, China 3. National Center for Rural Water Supply Technical Guidance, Chinese Center for Disease Control and Prevention, Beijing 102200, China

AR TIC LE I N FO

ABS TR ACT

Article history:

This study was performed to investigate the removal of phosphate from domestic wastewater

Received 16 July 2014

using a modified steel slag as the adsorbent. The adsorption effects of alkalinity, salt, water,

Revised 23 November 2014

and thermal modification were investigated. The results showed that thermal activation at

Accepted 1 December 2014

800°C for 1 hr was the optimum operation to improve the adsorption capacity. The adsorption

Available online 25 March 2015

process of the thermally modified slag was well described by the Elovich kinetic model and the Langmuir isotherm model. The maximum adsorption capacity calculated from the Langmuir

Keywords:

model reached 13.62 mg/g. Scanning electron microscopy indicated that the surface of the

Steel slag

modified slag was cracked and that the texture became loose after heating. The surface area

Thermal modification

and pore volume did not change after thermal modification. In the treatment of domestic

Phosphate

wastewater, the modified slag bed (35.5 kg) removed phosphate effectively and operated for

Adsorption

158 days until the effluent P rose above the limit concentration of 0.5 mg/L. The phosphate

Fractionation

fractionation method, which is often applied in soil research, was used to analyze the phosphate adsorption behavior in the slag bed. The analysis revealed that the total contents of various Ca–P forms accounted for 81.4%–91.1%, i.e., Ca10–P 50.6%–65.1%, Ca8–P 17.8%–25.0%, and Ca2–P 4.66%–9.20%. The forms of Al–P, Fe–P, and O–P accounted for only 8.9%–18.6%. The formation of Ca10–P precipitates was considered to be the main mechanism of phosphate removal in the thermally modified slag bed. © 2015 The Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences. Published by Elsevier B.V.

Introduction Phosphorus-rich wastewater discharge often causes eutrophication, which leads to the abundant growth of aquatic plants and algae. Therefore, domestic wastewater treatment plants are required to meet maximum P discharge limits. A control target of total phosphate (TP) for secondary effluent is between 0.5 and 1.0 mg P/L (Xiong et al., 2008). Accordingly, biological techniques are currently used in domestic waste⁎ Corresponding author. E-mail addresses: [email protected] (Jian Yu), [email protected] (Wenyan Liang).

water treatment, including activated sludge treatment (Coma et al., 2012), sequencing batch reactors (Jin et al., 2012), and biological contact oxidation (Lu et al., 2011). Biological treatments reduce most of the organic carbon and nitrogen in wastewater. The main obstacle is that they have been less effective in phosphorus removal. Therefore, these biological processes usually need to be combined with post-treatment procedures to remove phosphorus in advanced wastewater treatment. In the case of phosphorus removal, the adsorption technique has been widely studied in the past decades. Inorganic adsorbents, such as slag (Barca et al., 2012; Claveau-Mallet et al., 2012; Kostura et al., 2005), fly ash (Chen et al., 2007), red mud

http://dx.doi.org/10.1016/j.jes.2014.12.007 1001-0742/© 2015 The Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences. Published by Elsevier B.V.

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(Akay et al., 1997), alunite (Özacar, 2003), zeolite (Chen et al., 2006), soils (Gimsing et al., 2007; Moon et al., 2007), and iron oxide tailings (Zeng et al., 2004), have been widely used for phosphorus removal. Steel slag, a byproduct of the steel industry, has been considered a promising post-treatment option in terms of effluent quality, cost-efficiency, and operation simplicity. The application of steel slag as the adsorptive agent to remove phosphate from aqueous solution has been intensively attempted (Cucarella and Renman, 2009; Oguz, 2004). Batch experiments are commonly performed to evaluate the phosphate removal capacity of steel slag, which has been found to range from 1 to 80 mg/g (Bowden et al., 2009; Drizo et al., 2002; Jha et al., 2008; Xue et al., 2009). The main parameters affecting the efficiency of phosphate adsorption are the aggregate size, the ratio of material to solution, contact time, agitation, temperature, pH, and initial phosphate concentration (Cucarella and Renman, 2009). According to Xue et al. (2009), the anions Cl−, SO24 −, and NO−3 found in wastewater had insignificant effects on phosphate adsorption. Based on investigation of the effects of pH as well as chemical modeling, Baker et al. (1998) concluded that precipitation was the main mechanism of phosphate removal by slag treatment. In addition to precipitation, phosphate can also be removed by ion exchange and weak physical interactions between the surface of the sorbent and the metallic phosphate salts (Oguz, 2004). However, the use of slag in phosphate adsorption from aqueous solutions faces disadvantages, for example having a low specific surface area and a poor pore structure (Kostura et al., 2005). The internal chemical components cannot be used efficiently. Therefore, activating the steel slag surface is one effective way to utilize this waste. The methods of calcination, acidification, and basification have been applied to the slag to facilitate adsorption of ions + − 2+ such as PO3− 4 , F , Ni , and NH4, achieving an improvement of adsorption capacity (Islam and Patel, 2011; Jha et al., 2004, 2008; Li et al., 2011). Slag is mainly composed of calcium, iron, aluminum, and silicon oxides, similar to the composition of calcareous soils. Methods to analyze phosphate fractionation in calcareous soils are often used to understand the phosphate reaction mechanism when studying phosphate adsorption by sepiolite and palygorskite (Gan et al., 2009; Yin et al., 2011). According to the nature of the different cations that bind with phosphate, inorganic phosphate can be divided into calcium phosphate (Ca–P), iron phosphate (Fe–P), aluminum phosphate (Al–P), and occluded phosphate (O–P) (Jiang and Gu, 1989). Based on the numbers of combined phosphate groups, Ca–P can be further classified into dicalcium phosphate (Ca2(HPO4)2, Ca2–P), tricalcium phosphate (Ca3(PO4)2, Ca3–P), octacalcium phosphate (Ca8H2(PO4)6, Ca8–P), and ten-calcium phosphate (Ca10(PO4)6 (OH)2, Ca10–P) (Celen et al., 2007; Yin et al., 2011). Thus, in the present study, the steel slag was first activated to improve its adsorption capacity using alkaline, water, salt, or thermal modification. The surface properties of the modified slag were investigated. The phosphate adsorption isotherm and kinetics of the modified slag were evaluated in batch experiments. Then, a reactor filled with the modified slag followed the biological treatment in series to dispose of real domestic wastewater. The effectiveness of phosphate removal was evaluated. Phosphate fractions in different parts of the slag

bed were measured to quantify the amount of phosphate bound to mineral compounds in the slag and to understand the phosphate binding during treatment.

1. Materials and methods 1.1. Steel slag The steel slag was obtained from a steel plant in Jiangxi Province, China. Elemental analysis (expressed as oxides) using X-ray fluorescence spectroscopy (S4-Explorer, Bruker Co., Billerica, Massachusetts, Germany) showed that the proportional mass composition of the slag was Fe2O3 27.8%, CaO 49.75%, SiO2 10.5%, MnO 3.91%, MgO 2.28%, P2O5 1.80%, Al2O3 1.64%, V2O5 0.73%, TiO2 0.66%, Cr2O3 0.30%, SO3 0.28%, Na2O 0.25%, CuO 0.07%, and Nb2O5 0.03%.

1.2. Modification Before the experiments, the slag was sieved to yield the size fractions of 150–420 μm (mesh 100–40), 420–841 μm (mesh 40–20), 841–2000 μm (mesh 20–10), and 2000–4000 μm (mesh 10–5) using ASTM standard sieves, respectively. For thermal modification, the slag was heated directly at 800°C for 1 hr. For the alkaline, salt, or water modification, approximately 10 g of steel slag was immersed in 100 mL NaOH (3 mol/L), NaNO3 (1 mol/L), or distilled water for 24 hr. The screened slag was then heated at 800°C for 1 hr. In the evaluation of adsorption capacity, 0.5 g of slag was added to 100 mL of a PO3− 4 solution (100 mg P/L) prepared with KH2PO4. After the mixture was stirred at 160 r/min for 24 hr in a thermostatic shaker (25°C), the solution was centrifuged, and the filtrate was 3− used to measure the PO3− 4 concentration. The PO4 adsorption capacity at equilibrium was calculated using Eq. (1): qe ¼

ðC 0 −C e ÞV m

ð1Þ

where, C0 (mg P/L) and Ce (mg P/L) are the initial and equilibrium concentrations of PO3− 4 , m (g) is the dry mass of the slag, and V (L) is the solution volume.

1.3. Reaction devices As shown in Fig. 1, the experimental system included two parts, namely, biological contact oxidation (BCO) and steel slag adsorption (SSA). The BCO and SSA reactors, with working volumes of 20 and 40 L, respectively, were fabricated from polymethyl methacrylate. Volcanic stones and ceramic rings were chosen as the biofilter media due to their specific area and low cost. The adsorption bed in the SSA reactor was composed of two layers: 5 cm of gravel and 30 cm of the thermally modified slag (35.5 kg, size 2000–4000 μm).

1.4. Reactor operation The raw wastewater discharge from the student dormitories was used, with the characteristics of pH 6.8–7.6, chemical oxygen demand (COD) 129.5–308.7 mg/L, ammonia nitrogen

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(NH+4-N) 6.5–26.5 mg/L, total nitrogen (TN) 11.8–28.9 mg/L, and TP 1.0–2.7 mg P/L. After the start-up of the BCO reactor (see SI), the raw wastewater was pumped into the reactor at a flow rate of 50 mL/min. Air was supplied through a diffuser at 3.3 L/min to keep the dissolved oxygen (DO) concentration in the range 4–5 mg/L. The effluent from the BCO reactor was collected in a storage tank and then continuously pumped into the SSA reactor with a peristaltic pump at a flow rate of 7.2 mL/min. Influent and effluent samples of both reactors were collected every three or four days to measure the concentrations of COD, NH+4-N, TN, and TP.

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Ca8–P); (3) 0.5 mol/L NH4F (pH = 8.2), 20–25°C for 1 hr (to obtain Al–P); (4) 0.1 mol/L NaOH–0.05 mol/L Na2CO3, 20–25°C for 4 hr (to obtain Fe–P); (5) 0.3 mol/L Na3C6H5O7·2H2O–Na2S2O4 (1 g)– 0.5 mol/L NaOH, 80–90°C, for 25 min (to obtain O–P); and (6) 0.25 mol/L H2SO4, 20–25°C for 1 hr (to obtain Ca10–P). During every step of the extraction, the samples were centrifuged, and the supernatant was analyzed with the ascorbic acid method. All of the experiments were conducted in triplicate, and the data were expressed using the average values.

2. Results and discussion 1.5. Analytical determinations 2.1. Slag modification Analyses of TP were conducted with an ultraviolet spectrophotometer using the ascorbic acid method (Yang et al., 2009). The NH+4-N and TN concentrations were measured according to the National Standard Methods (Yang et al., 2009). The COD was measured with a COD analyzer (CTL-12, Chengde Huatong Environmental Protection Apparatus Co., Chengde, Hebei Province, China). The DO was continuously measured online using a DO detector (Multi 3410, WTW GmbH, Weilheim, Germany). The scanning electron microscopy was performed using an S-3400N microscope (Hitachi Co., Chiyoda, Tokyo, Japan) operated at 10 kV. The X-ray diffraction (XRD) pattern was obtained using a D/MAX 2000 (Rigaku Co., Tokyo, Japan) with monochromatic Cu Kα radiation. The surface areas calculated by the Brunauer–Emmett–Teller (BET) equation were determined from N2 adsorption isotherms measured with a NOVA4000 system (Quantachrome Co., Boynton Beach, Florida, USA). The sample was degassed under vacuum at 200°C for 11 hr prior to data collection. The pore size distribution was calculated using the Barret–Joyner–Halenda algorithm.

1.6. Phosphate fractionation of the adsorbed phosphorus After the SSA reactor stopped running, slag from the reactor was sampled to examine the phosphate fractions adsorbed on it, which were extracted as discrete chemical fractions by a series of increasingly strong reagents to obtain more recalcitrant forms (Yin et al., 2011). In brief, the fractionation procedure was performed as follows (Gan et al., 2009): 1.000 g air-dried samples were transferred to 100 mL flasks and sequentially extracted with (1) 0.25 mol/L NaHCO3 (pH = 7.5), 20–25°C for 1 hr (to obtain Ca2–P); (2) 0.5 mol/L NH4Ac (pH = 4.2), 20–25°C for 1 hr (to obtain

The results of the modification are presented in Fig. 2. Compared with the unmodified slag, the adsorption capacities for the size of 2000–4000 μm increased by 191%, 107%, 125%, and 118% after alkali, salt, water, and thermal modification, respectively. It was clear that the alkali-modified slag possessed the highest adsorption capacities among the four methods. The greater the alkalinity, the higher the activity of the slag and the higher the adsorption of phosphate (Jha et al., 2008). The use of sodium hydroxide in this study provided a certain degree of basicity to the modified slag, and therefore gained more adsorption capacity than those in other methods. Other than the alkaline modification, the other three methods (i.e., salt, water, and thermal modification) displayed insignificant differences in adsorption capacity. This indicated that the heating step was actually the key factor that determines the modification effects. The increase of adsorption capacity after water and salt modification was mainly a result of heat treatment and not of the water or salt soaking. The subsequent investigations (Fig. S1) showed that 800°C and 1 hr were the optimal temperature and time parameters for the thermal modification. Although the alkaline modification had the highest phosphate adsorption, the alkaline reagents increased the cost of the slag modification. Therefore, thermal modification was considered to be the most appropriate method for this study. To further illustrate the effects of thermal modification, a small column was filled with the modified slag to perform dynamic experiments. The slag particle size of 2000–4000 μm was chosen in the experiments to avoid clogging. The results in Fig. S2 show that the column filled with modified slag could run for 73.7 hr before the effluent phosphate reached 0.5 mg

Fig. 1 – Schematic diagram of the system for phosphate removal from domestic wastewater using biological contact oxidation and the modified steel slag adsorption.

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P/L, while the column with untreated slag ran for only 14.8 hr. The thermal modification successfully ameliorated the phosphate adsorption capacity of the slag in the column tests. The increment of adsorption capacity was less obvious as the size decreased. Compared with the unmodified slag, the adsorption capacities for the size of 150–420 μm were only raised by 17.9%, 9.03%, 6.82%, and 3.86% for the alkali, salt, water, and thermal modifications, respectively. The specific surface area of the slag grew as the particle size was reduced, increasing the adsorption capacity and reducing the relative effects of the modifications. As seen in Fig. 3, the slag surface cracked and its texture became loose after thermal modification. However, the surface area of the thermally modified slag (in Table 1) increased only 1.70%. The uncalcined slag mainly contained larnite (β-Ca2SiO4) and plustite (FeO) in the study of Jha et al. (2004). After calcining, kirschsteinite (CaFeSiO4) and manganese oxide (Mn3O4) appeared as a result of solid-state reaction between iron oxide and calcium silicate. Thus, the improvement of adsorption capacity for thermal modification was not caused by the change of the surface area but by the structure and mineral composition of the slag. It also can be seen from Fig. 3c that some particles adsorption. The precipitated on the slag surface after PO3− 4 slag surface became fine and closed again after the surface substances reacted with the phosphate.

2.2. Kinetics of phosphate adsorption A kinetic study on PO3− 4 adsorption was carried out on the thermally modified slag at the size of 2000–4000 μm. As shown in Fig. S3, the time to reach adsorption equilibrium was nearly the same before and after modification. The majority of PO3− 4 adsorption on the slag was completed in 24 hr. To quantify the changes in adsorption as a function of time, the adsorption kinetic data were fitted with several kinetic models (i.e., pseudo-

first order, pseudo-second order, and Elovich equations). The equations were rearranged in their linear form as follows:  Pseudo‐first order

Pseudo‐second order

Elovich

qt ¼

ln ðqe −qÞ ¼ ln qe −K 1 t 1 t
þ ¼ qt qe K 2 q2e

1 1 ln ðαβÞ þ lnt β β

 t 2:303

ð2Þ

ð3Þ

ð4Þ

where, qe (mg/g) and qt (mg/g) are the PO3− 4 amount adsorbed at equilibrium and at time t (hr); and K1, K2, α and β are rate constants of Eqs. (2)–(4), respectively. The pseudo first-order (and also the pseudo second-order) model predicts the behavior over the “whole” range of studies, supporting its validity. Moreover, it also shows that chemisorption is the rate-controlling factor (Gan et al., 2009). The Elovich equation does not predict any definite mechanism, but it is useful in describing adsorption on highly heterogeneous adsorbents (Zeng et al., 2004). As shown in Table 2, the adsorption data of untreated slag could be well fitted to these models. It was observed that the correlation coefficient (R2 = 0.992) of the pseudo-first-order kinetic model was a little higher compared to the other two kinetic models, suggesting that the overall adsorption rate of the PO3− 4 should be controlled by the chemical process in the pseudo-first order reaction mechanism. When experimental data fit well with the pseudo-first order model, the chemical sorption system involves valence forces through sharing or exchanging of electrons between adsorbent and adsorbate (Gan et al., 2009). After thermal modification, the highest correlation coefficient (R2 = 0.978) was obtained with the Elovich equation, indicating that heterogeneous adsorption was more appropriate for the modified slag. This high applicability of the Elovich equation is generally in agreement with other researchers' results, where the Elovich equation was able to properly describe the kinetics of phosphate adsorption on soils and soil minerals (Zeng et al., 2004).

2.3. Adsorption isotherms

Fig. 2 – Effects of modification on the adsorption capacity of phosphate using steel slag with different particle size. The four modifications are under the calcinations at 800°C for 1 hr followed by alkaline, salt, water, and thermal modification, respectively.

An adsorption isotherm is often used to describe how molecules of the sorbate interact with the adsorbent surface. Thus, PO3− 4 uptake experiments were carried out at different initial concentrations ranging from 10–125 mg P/L. The data are presented in Fig. S4. The uptake of PO3− 4 initially increased sharply with increasing initial concentration and finally slowly reached saturation. A large amount of released Ca2+ increases the solution pH and favors the precipitation of the calcium-phosphate compound (Jha et al., 2008). Therefore, resulted in increasing higher initial concentration of PO3− 4 calcium phosphate precipitation, which decreased the pH values. The adsorption of PO3− 4 ions on the slag surface occurs at relatively low pH values. As the PO3− 4 is adsorbed at the slag surface, the surface becomes saturated for further interactions with other molecules. The phosphate adsorption behavior can be simulated by the use of mathematical equations. Langmuir and Freundlich models were used to simulate the isotherm data in this work. The Langmuir isotherm assumes that the uptake of ion occurs

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Fig. 3 – Scanning electron microscope patterns of steel slag with the particle size of 2000–4000 μm (mesh 10–5). (a) Before modification; (b) after thermal modification; (c) thermally modified slag after phosphate adsorption. on a homogenous surface by monolayer sorption due to surface saturation being the limiting factor. The Freundlich isotherm assumes that the uptake of ion occurs on a heterogeneous surface by multilayer adsorption (Jha et al., 2008). The models take the following forms: Langmuir

Ce Ce 1 ¼ þ qe qm Kqm

ð5Þ

Freundlich

qe ¼ K F C 1=n e

ð6Þ

where, Ce (mg P/L) is the equilibrium concentration, qe (mg/g) is the amount adsorbed at equilibrium, qm (mg/g) and K (L/mg) are the Langmuir constants, and KF and n are the Freundlich constants. The parameters calculated for the Langmuir and Freundlich equations are listed in Table 3. The equilibrium data were fitted better with the Langmuir model than with the Freundlich model for both the modified and unmodified slag. This finding suggested surface homogeneity for the adsorbent and monolayer adsorption onto a surface of identical sites. Similar results have been reported for the adsorption of phosphate on untreated slag, Al(OH)3-activated slag, alunite, and slag–coal cinder (Kostura et al., 2005; Özacar, 2003; Yang et al., 2009). The values of maximum adsorption capacity (qm) suggested that the modified slag (qm = 13.620) behaved better than the unmodified one (qm = 8.560). This was due to the loose structure of the modified slag, which made the Ca2 + more easily released into the solution to react with PO34 −. The qm parameters of the modified slag could be compared with other results from the literature. Xiong et al. (2008) investigated P sorption capacities of steel slag through magnetic separation and obtained a qm of 5.3 mg/g. The sorption capacities of blast furnace slag varied from 0.65 to 44.2 mg/g even with very similar chemical compositions (Kostura et al., 2005). Because there are many factors affecting the P adsorption capacities of the materials used in batch studies, the phosphorus adsorption

Table 1 – Surface measurements of the slag before and after thermal modification. SBET Pore volume Mean pore (m2/g) (m3/g) size (nm) Before modification After thermal modification

7.523 8.622

0.011 0.012

5.912 5.557

SBET: surface area calculated using the Brunauer–Emmett–Teller model.

capacities of derived materials can vary by several orders of magnitude.

2.4. Phosphate removal in long-term operation The BCO reactor was operated and its performance was observed regularly. The results of COD, NH+4-N, TN, and TP during the start-up phase are presented in Fig. S5. After approximately 20 days, the surface of the filter media was covered with a slimy, brown biofilm, while the treatment processes remained stable. At this point, it was determined that the start-up phase of BCO was finished. After the successful start-up of the BCO reactor, the SSA process was connected to the system and followed the biological treatment. The results of COD, NH+4-N, and TN are presented in Fig. S6. During most of the long run, their concentrations in the BCO system met the discharge standard of municipal wastewater treatment plants in China. However, the TP in the BCO effluent did not satisfy water quality standards (0.5 mg P/L). In the biological treatment to remove phosphorus, the anoxic phosphorus-accumulating bacteria adsorb excess compounds from the environment. They store polyphosphate and form a sludge with a high phosphorus content, which can be removed by sludge discharge (Hocaoglu et al., 2011). The production of sludge in the BCO reactor was so small that there was barely any sludge to discharge. Moreover, the removal process of the system lacked the necessary anaerobic or aerobic circulation, so only a fraction of phosphorus was taken up by the growth of microorganisms. Most of the phosphorus was discharged with the biological effluent. The principal forms of phosphorus in human wastewater are organically bound phosphorus, polyphosphates, and orthophosphates. Organically bound phosphorus and polyphosphates can be hydrolyzed and converted to orthophosphates during the aerobic biological treatment (Yeoman et al., 1986). Thus, the principal form of phosphorus in the effluent of the BCO system is assumed to be orthophosphates, although the other forms may exist. When the BCO process was combined with the SSA, the TP removal efficiency increased obviously as shown in Fig. 4. In the initial 60 days, the TP concentration of the SSA effluent was under the detection limits (0.01 mg P/L), indicating that the phosphate in the BCO effluent was removed by the SSA processes. Overtime, the TP removal efficiency decreased gradually. After 158 days of operation, the TP concentration of the SSA effluent exceeded 0.5 mg P/L and the reactor was stopped. Although the SSA removed phosphate effectively,

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Table 2 – Estimated kinetic model constants for phosphate adsorption on the slag. Temperature (25°C)

First-order equation −1

Slag without modification Thermally modified slag

Second-order equation 2

Elovich equation 2

K1 (hr )

qe (mg/g)

R

K2 (g/(mg · hr))

qe (mg/g)

R

α

β

R2

0.240 0.299

12.66 17.58

0.992 0.928

0.006 0.011

15.53 19.71

0.986 0.962

3.79 10.23

0.33 0.26

0.974 0.978

Table 3 – Estimated isotherm model constants for phosphate adsorption on the slag. Temperature (25°C)

Slag without modification Thermally modified slag

Langmuir constants

Freundlich constants

qm (mg/g)

K (L/mg)

K·qm (L/g)

R2

KF (g·(mg/L))

1/n

R2

8.560 13.620

0.575 0.610

4.922 8.308

0.999 0.997

4.973 7.541

0.131 0.139

0.660 0.616

the concentrations of COD, NH+4-N, and TN in the SSA effluent were close to their values in the BCO effluent. The slag bed had almost no effects on their removal. Organic chemicals and nitrogen compounds did not influence the operation of the SSA in terms of phosphate removal.

2.5. Phosphate fractionation in the slag bed When wastewater flowed through the slag bed, the PO3− 4 in the wastewater adsorbed on the surface of the slag and reacted with the cations on the slag, generating a variety of phosphate-cation precipitates (Baker et al., 1998). Therefore, a phosphate fractionation scheme was performed on the P-adsorbed slag. After 158 days of operation, the SSA reactor was stopped and air-dried for approximately 20 days. Then, the slag samples were taken out from different parts of the reactor, as shown in Fig. 5a. As illustrated in Fig. 5b, the Ca–P forms (sum of Ca2–P, Ca8–P, and Ca10–P) accounted for most of the phosphate fractionation (81.4%–91.1%), whereas the sum of Al–P, Fe–P, and O–P accounted for only 8.9%–18.6%. The relatively high percentage of Ca–P indicated that the high concentration of phosphate removed by the modified slag was effected mainly by calcium phosphate precipitation. In the Ca–P fractionation, the Ca10–P content (50.6%–65.1%) was clearly higher than other calcium-bound forms. The Ca8–P content (17.8%–25.0%) was much lower than that of Ca10–P, with the Ca2–P content being the smallest (4.66%– 9.20%). During the formation of calcium phosphate, the phosphates in solution combine with the calcium in the media to generate Ca2–P, Ca8–P, and Ca10–P, according to the following equations (Yin et al., 2011): Ca2þ þ 2H2 PO4 − þ H2 O→CaðH2 PO4 Þ2  H2 O

ð7Þ

CaðH2 PO4 Þ2 þ Ca2þ →2CaHPO4 þ 2Hþ

ð8Þ

6CaHPO4 þ 2Ca2þ →Ca8 H2 ðPO4 Þ6 þ 4Hþ

ð9Þ

Ca8 ðH2 PO4 Þ6 þ 2Ca2þ þ H2 O→Ca10 ðPO4 Þ6 ðOHÞ2 þ 4Hþ

ð10Þ

It is known that the pKa1, pKa2, and pKa3 of phosphoric acid are 2.15, 7.20, and 12.35, respectively. Between pH of 3 and 7, the predominant species were H2PO−4 ions, and the order

of Ca–P content was distributed as Ca8–P (75.2%) > Ca2–P (18.1%) > Ca10–P (2.50%) in the Ca–P fractionation of P-saturated sepiolite (Yin et al., 2011). Large amounts of released Ca2 + from the slag increased the pH of the solution and led the pH values of the SSA effluent to be 12.89–9.78 in the running stage. The predominant species in that range are mainly HPO24 − and PO34 −. Thus, the strongly alkaline environment favors the production of Ca10–P. In addition to the pH, the kinetics of Ca8–P and Ca10–P production is extremely slow, and it would take a few days to form Ca8–P and Ca10–P (Celen et al., 2007). When the wastewater flowed from the bottom to the top layer of the slag bed, the hydraulic duration time of phosphate in the reactor was 3.86 days, which made the Ca10–P content increase obviously with the flow direction and the altitude of the bed. In addition to Ca–P fractionation, other phosphate fractions occupied a small percentage. The Al–P content was approximately 1.83%–5.48%, close to the mass fraction of aluminum oxides in the slag (1.64%). However, the Fe–P content was only approximately 3.97%–8.00%, which was less than the mass fraction of iron oxides in slag (27.84%). In wastewater applications, the most common and successful method to precipitate

Fig. 4 – Variation of phosphate concentration and the removal efficiency during the processes of biological contact oxidation (BCO) and steel slag adsorption (SSA). Removal efficiency refers to the total phosphate removal of the BCO and SSA reactors.

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Fig. 5 – Parts of the slag bed where phosphate fractionation was measured (a) and phosphate fractionation of the adsorbed phosphorus at different locations of the slag bed (b). U: upper layer; M: middle layer; B: bottom layer.

3+ 2+ 3+ PO3− and Fe2+ 4 involves dissolved cations such as Al , Ca , Fe (Yin et al., 2011). Because the phosphate adsorption by iron oxides is favored under low pH conditions (Li et al., 2013; Sousa et al., 2012), the strong alkaline pH of the slag led to weak reaction of phosphate with iron oxides. The occluded phosphates are those that are wrapped in an insoluble film of calcium phosphate, iron phosphate, aluminum phosphate, and other phosphate forms. They form a class of highly insoluble inorganic phosphates. In particular, the iron oxides form an inert film wrapped on the surface of the phosphate, which greatly reduces the solubility of the original phosphate (Sousa et al., 2012). Therefore, the O–P did not readily participate in the chemical reactions unless the outer film was removed. In this study, the content of the O–P fraction was approximately 2.80%–6.86%. Above all, the adsorption capacities of Ca2–P, Ca8–P, and Ca10–P for modified slag were clearly higher than those of the untreated slag (Fig. S7); whereas, the adsorption capacities of Al–P, Fe–P, and O–P of the modified slag had little difference with those of the untreated slag. The results illustrate that the increment of adsorption effects of the modified slag comes from the reaction between calcium and phosphate. There was a white layer of precipitate in the top of the reactor. The white precipitates were small and light and could flow with the wastewater out of the bottom and middle area in to the upper layers. The XRD spectra in Fig. S8 showed sharp peaks, indicating that the precipitates were crystalline in nature. The composition of the precipitates corresponded to the theoretical composition of calcium carbonate. The XRD spectra showed no signals from the Ca–P, Al–P, and Fe–P forms. This indicated that the Ca–P, Al–P, Fe– P, and O–P were mainly adsorbed on the slag surface through monolayer adsorption, which showed the character of a Langmuir isotherm. These phosphate fractions adsorbed on the slag in the wastewater from reacting surface hindered the PO3− 4 further with the slag and stopped the reactor.

3. Conclusions This study showed that thermal modification of steel slag was the most appropriate and cost-effective method to improve

the adsorption capacity for phosphate. The phosphate adsorption from aqueous solutions by thermally modified slag was found to follow Elovich type kinetics and a Langmuir adsorption isotherm. It was proved that the reactor filled with the thermally modified slag removed phosphate efficiently from the biological effluent. The P-adsorbed slag bed showed the phosphate fractionation order of Ca10–P (50.6%–65.1%) > Ca8–P (17.8%–25.0%) > Ca2–P (4.66%–9.20%) > Fe–P (3.97%–8.00%) > O–P (2.80%–6.86%) > Al–P (1.83%–5.48%). A greater amount of Ca10–P accumulated at the top area of the reactor. The Ca2–P, Ca8–P, Al–P, Fe–P, and O–P were distributed evenly throughout the slag bed. The phosphate fractionation revealed that the main mechanism of phosphate removal was the chemical reaction of phosphate with calcium in the slag and the formation of Ca10–P precipitates.

Acknowledgments This work was supported by the Mega-projects of Science Research for Water Environment Improvement of China (Nos. 2013ZX07209-001-003, 2012ZX07307-001-006).

Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx.doi.org/10.1016/j.jes.2014.12.007.

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