Flue gas desulfurization (FGD) wastewater treatment by conventional neutralization, chemical precipitation and coagulation process removes most suspended solids and heavy metals, and provides an effluent rich in calcium, alkalinity and chloride, which obstructs its reclamation and reuse but is in favor of phosphorus (P) precipitation. The goals of this study were to investigate feasibility of reusing FGD effluent as a calcium source for P removal from P-rich wastewater. Results revealed that increasing the volumetric ratio between FGD effluent and P-rich wastewater achieved higher pH value and Ca/P ratio, and thus enhanced P removal efficiency to 94.3% at the ratio of 40%. X-ray diffraction and scanning electron microscope analysis of harvested precipitates showed that increasing pH from 8 to 10 induced the conversion of hydroxyapatite to tri-calcium phosphate, and then to whitlockite. This study demonstrated that for reusing FGD effluent for P removal was highly feasible, both technically and economically. This process not only saves the cost of precipitants for P removal, but also provides an economical alternative for current zero liquid discharge technology for FGD wastewater, which requires high energy consumption and capital costs.

INTRODUCTION

Coal-fired power plants currently supply 41.3% of global electricity and, in some countries, coal fuels a higher percentage of electricity (International Energy Agency 2015). In China, coal-fired power generation capacity was 3.95 trillion kWh, accounting for 70.5% of total electrical generating capacity in 2014 (China Electricity Council 2014). Because of the vast amount of atmospheric pollutants emission, flue gas desulfurization (FGD) process is commonly used for sulfur dioxide removal from the exhaust flue gas (Guan et al. 2009; Huang et al. 2013a). Among various FGD technologies, the lime/limestone–gypsum process is the most commonly installed scrubber with advantages of high desulfurization efficiency, high reliability, high adaptability and low cost (Deng et al. 2014). For the corrosion protection of the equipment and gypsum quality requirements, a certain amount of wastewater was discharged periodically from the FGD system to maintain concentrations of chloride and heavy metals less than the designed level (Guan et al. 2009; Deng et al. 2014). The discharged FGD wastewater has low pH, high concentrations of suspended solids (SS), chloride and hardness, and is abundant in heavy metals (Guan et al. 2009; Huang et al. 2013a, 2013b, 2013c).

Conventional FGD wastewater treatment process, which consists of neutralization, chemical precipitation and coagulation (NPC) units, is extensively applied to neutralize pH, precipitate heavy metals and remove SS. NPC process removes most SS and heavy metals, and provides an effluent still rich in calcium, sulfate and chloride, which no longer met either new stringent discharge limits, or water conservation and reuse requirements, such as Water Pollution Control Action Plan in China (State Council-2015-17), Effluent Limitations Guidelines and Standards for the Steam Electric Power Generating Point Source Category in the USA (EPA-HQ-OW-2009-0819-5558), etc. Zero liquid discharge (ZLD) process, consisting of evaporation and crystallization, has been developed to solve the new challenge (Shaw 2008; Deng et al. 2014), and to reuse the treated FGD wastewater in power plants. However, high capital cost and energy consumption are the main obstacles for application of the ZLD process (Pakzadeh et al. 2014). As a sulfur source, FGD wastewater has been successfully utilized to achieve beneficial co-treatment with municipal wastewater through a sulfate reduction, autotrophic denitrification and nitrification integrated process (Qian et al. 2015). Therefore, reusing the FGD effluent for resources recovery is a potential sustainable way.

The FGD effluent in the conventional NPC process, before being neutralized by acid, contained high Ca2+ and alkaline, and was reused for phosphorus (P) removal from P-rich wastewater for the first time in this study. P-rich wastewater usually generates from phosphate fertilizer production, piggery, sludge dewatering and anaerobic digestion, etc., with P concentrations ranging from 50 to 200 mg/L (Huang et al. 2014; Hu et al. 2015; Ren et al. 2015). Effects of pH and dosage of FGD wastewater added into the P-rich wastewater (expressed as their volumetric ratio, FPR) on P removal efficiency (PRE) were investigated. The collected precipitates at different initial pH values were analyzed by scanning electron microscope (SEM) and X-ray diffraction (XRD) to evaluate their compositions and morphology. The results are expected to provide useful information on reusing FGD effluent as a low-cost alternative way for P removal from wastewater.

METHODS

Wastewater source

FGD wastewater used in this study was taken from Changshu Power Plant, (Jiangsu, China), and its characteristics are shown in Table 1. FGD wastewater was treated according to the following steps to obtain effluent of conventional NPC process. (1) Neutralization: FGD wastewater was neutralized by the addition of CaO to pH of 9.0–9.5, and then let stand for 30 min. In this process, most heavy metals and fluoride were removed. (2) Chemical precipitation: the obtained supernatant was precipitated by 25 mg/L Na2S, stirred for 3 min at 300 rpm, and then let stand for 30 min. Heavy metals were reduced to meet the relevant discharge standards (e.g. Integrated Wastewater Discharge Standard of China, GB8978-1996) in this process. (3) Coagulation: the precipitated supernatant was coagulated with 60 mg/L polymeric aluminum and 1 mg/L cationic polyacrylamide to remove SS, stirred for 10 min at 100 rpm, and then left to stand for 30 min.

Table 1

Characteristics of raw FGD wastewater and effluent of NPC process

IndexRawEffluent
pH 6.87 9.11
SS (g/L) 86.6 0.39
COD (mg/L) 255 115
Ca2+ (mmol/L) 34.2 149.6
Mg2+ (mmol/L) 193.8 193.8
Cl (g/L) 33.9 25.9
Hg (μg/L) 0.271 0.042
Cd (mg/L) 0.391 0.007
Cr (mg/L) 0.006 0.003
NH4+-N (mg/L) 0.23 0.12
PO43−-P (mg/L) 0.02 <0.01
TP (mg/L) 0.16 0.08
IndexRawEffluent
pH 6.87 9.11
SS (g/L) 86.6 0.39
COD (mg/L) 255 115
Ca2+ (mmol/L) 34.2 149.6
Mg2+ (mmol/L) 193.8 193.8
Cl (g/L) 33.9 25.9
Hg (μg/L) 0.271 0.042
Cd (mg/L) 0.391 0.007
Cr (mg/L) 0.006 0.003
NH4+-N (mg/L) 0.23 0.12
PO43−-P (mg/L) 0.02 <0.01
TP (mg/L) 0.16 0.08

Preparation of P-rich wastewater

In order to simulate the P-rich wastewater with P concentrations ranging from 50 to 200 mg/L, artificial wastewater was prepared by dissolving an analytical reagent grade of Na2HPO4·6H2O chemicals in distilled water to get an initial orthophosphate phosphorus (PO43−-P) concentration of 163.7 mg/L.

Batch tests and sampling

The treated FGD wastewater was used for P removal from P-rich wastewater. In the experiment to analyze effects of pH on PRE and precipitates composition, the initial pH values were regulated from 7.0 to 10.5 with an interval of 0.5, and FPR was maintained at 25%. To investigate the effect of FPR on PRE, volumes of FGD added into 200 mL P-rich wastewater were adjusted to achieve FPRs at 5%, 10%, 20%, 30%, 40%, 50% and 60%. FGD wastewater and P-rich wastewater was mixed for 5 min at 100 rpm by magnetic stirring in all the experiments. Then the mixed wastewater was settled for 30 min, and samples were collected from beakers for subsequent analysis. The precipitates collected at initial pH values of 8.0, 9.0 and 10.0 were purged with deionized water and dried for the analysis of composition and morphology. All the experiments were performed in duplicates at 25 ± 1 °C.

Analytical method

Chemical oxygen demand (COD), SS, ammonium nitrogen (NH4+-N), total phosphorus (TP) and PO43−-P were measured according to standard methods (Chinese NEPA 2012). The pH values were continuously measured and recorded by HQ30d (Hach, USA). Concentrations of Ca, Mg, Cd and Cr were determined using a 7510 inductively coupled plasma – atomic emission spectroscopy (Shimadzu, Japan). The concentration of Hg was determined using a RA-915+ portable mercury analyzer (Lumex Ltd, Russia). The concentration of chloride was measured by Mohr method. The morphology of crystals was observed with S-4800 SEM (Hitachi, Japan). The precipitates were characterized by D8 Advance Powder XRD (40 kV, 40 mA, step size 0.1°, Bruker Ltd, Germany). The significance of factors on PRE was analyzed by one-way factor analysis of variance using Office Excel 2007 (Microsoft, USA).

RESULTS AND DISCUSSION

Treatment performance of NPC process

As shown in Table 1, raw FGD wastewater contained high concentrations of SS, Ca2+, Mg2+ and chloride, a certain amount of COD and heavy metals, and negligible nutrients. After the NPC treatment, concentrations of SS, COD and heavy metals decreased greatly, and pH value and the concentration of Ca2+ were enhanced owing to the addition of CaO (Table 1). Considering the calcium phosphate precipitated in the form of thermodynamically stable hydroxyapatite (Ca5(PO4)3(OH), HAP), the FGD effluent completely removed P from P-rich wastewater at FPR of 6% in theoretical. COD in the FGD effluent is usually composed of thiosulfate, sulfite and rhodanate, which can be easily removed by aeration or the addition of oxidants before used for P removal. Nevertheless, P removal by the precipitation of calcium phosphate hardly occurred when pH approached to neutral condition, because the effective pH range for calcium phosphate crystallization was 8.5–9.0 (Qiu et al. 2012, 2015). Saxena & Bassi (2013) reported that phosphorus was effectively removed by strong alkali treatment from hydroponic greenhouse effluent abundant in calcium. In this study, high pH of the FGD effluent favored P removal from P-rich wastewater without addition of external alkaline substances.

Effect of pH on PRE

Qiu et al. (2012) reported that higher OH promotes crystallization of calcium phosphate. At FPR of 25%, Ca2+ and PO43−-P in the mixed wastewater were 1,197 and 128 mg/L, respectively. Figure 1(a) shows that the PRE rose from 68.1% to 96.2% when initial pH increased from 7.0 to 8.5, and then remained above 97.2% at initial pH of 9.0–10.0. One-way factor analysis of variance demonstrated that initial pH had significant effects on PRE at p < 0.01 level. These results indicated that the precipitation of calcium phosphate occurred under neutral condition, and the PRE maintained high and stable at pH > 8.5. High alkalinity was the dominating factor in calcium phosphate crystallization. Generally, at pH > 7, the likely precursor phase is an amorphous calcium phosphate (Cax(PO4)y·nH2O), followed by recrystallization into thermodynamically more stable HAP with the increase of pH (Valsami-Jones 2001). With pH increasing from 7.0 to 9.0, the thermodynamic driving force for the precipitation of calcium phosphate strengthened (Song et al. 2002), and thus PRE raised greatly.
Figure 1

Effect of pH on removal of PO43−-P, Ca2+ and Mg2+ by reusing FGD effluent.

Figure 1

Effect of pH on removal of PO43−-P, Ca2+ and Mg2+ by reusing FGD effluent.

Figure 1(b) presents concentrations of calcium and magnesium under various pH values. As pH rose, the concentration of calcium decreased in a high coincidence with the increase of PRE. A drastic decline of calcium concentration was observed with pH increasing from 7 to 9, and then the calcium concentration remained unchanged with pH increasing. It was also indicated that alkaline condition was beneficial for the formation of calcium phosphate precipitates. Furthermore, FGD wastewater contained considerable Mg2+, which competed with Ca2+ to react with phosphate and had adverse effects on structure of precipitates. Ferguson & McCarty (1971) reported that Mg2+ kinetically hindered nucleation and subsequent growth of HAP by competing for structural sites with chemically similar but larger Ca2+. As the pH increased from 7.0 to 10.0, the concentration of Mg2+ reduced by 4%, resulting in a small amount of Mg2+ incorporated into the structure of calcium phosphate precipitation. Thus, the higher alkalinity was conducive to change the composition of precipitates, while the interaction between Mg2+ and pH on phosphorus removal was negligible.

XRD and SEM analysis of the harvested precipitates

In order to evaluate effects of pH on the formation of precipitates, the composition and morphology of the precipitated crystals were characterized by XRD and SEM analysis. In Figure 2(a), each pattern included HAP as the dominant phase, with tri-calcium phosphate (Ca3(PO4)2·nH2O, TCP) and calcium sulfate hydrate (CaSO4·0.15H2O) as the minor phase at pH 8. However, the main phase was converted to TCP at pH 9 (Figure 2(b)) and to whitlockite ((Ca,Mg)3(PO4)2) at pH 10 (Figure 2(c)), respectively.
Figure 2

XRD patterns of precipitates harvested from P removal by reusing FGD effluent. The peaks are labeled T, TCP; H, HAP; C, CaSO4·0.15H2O and W, whitlockite.

Figure 2

XRD patterns of precipitates harvested from P removal by reusing FGD effluent. The peaks are labeled T, TCP; H, HAP; C, CaSO4·0.15H2O and W, whitlockite.

With Jade 6.5 software to analyze the XRD pattern, one well-crystallized HAP-derived main peak appeared at 31.8°, and corresponded to Miller indices (211). The poor crystalline was marked by two TCP peaks at 2θ of 21.8° and 45.5° (Okano et al. 2015) and three calcium sulfate hydrate peaks at 2θ= 25.5°, 29.6° and 49.3°. The formation of HAP under a neutral condition was also observed by Zhou et al. (2012) for phosphate removal on hydrocalumite. The coexistence of HAP and TCP was also observed at pH 9 with TCP as the major component, with the main peak appearing at 31.9°. The increase of pH from 8 to 9 resulted in a significant inhibitory effect on the conversion of TCP to HAP, which was in good agreement with the findings in this study. High alkalinity favored formation of thermodynamically unstable TCP because energy savings in the precipitation greatly outweighed those of lattice ordering into HAP (van der Houwen & Valsami-Jones 2001). At pH of 10, whitlockite was the major component with peaks occurring at 31.5° and 45.0°, corresponding to Miller indices (217) and (324). The formation of whitlockite indicated the incorporation of Mg2+ into calcium phosphate. Furthermore, the presence of Mg2+ had great inhibitory effects on HAP growth, and the coprecipitation of Mg2+ with calcium phosphate played a facilitative role in the formation of amorphous calcium phosphate instead of HAP (Suchanek et al. 2004). Cao & Harris (2007) also pointed out Mg2+ incorporated into the structure of HAP, causing structural changes that inhibit the formation of HAP. The reported results are well confirmed by findings in this study.

Figure 3 illustrates the SEM images of harvested precipitates at different pH values. The precipitates had smooth surface and agglomerated as small irregular particles at pH 8 (Figure 3(a)), which was the typical morphology of HAP (Eliassi et al. 2014). Figure 3(b) displays an image of TCP, which demonstrated the confluence of spherulitic particles with the observed particle sizes about 0.01–0.1 μm (Ren et al. 2015). In Figure 3(c), the morphology of the crystals was similar to those in Figure 3(b), but more compact, which was caused by the incorporation of a small percentage of Mg2+ that changed the structure and converted the precipitates to whitlockite.
Figure 3

SEM analysis of precipitates harvested from P removal by reusing FGD effluent at different pH values magnified at 100 k.

Figure 3

SEM analysis of precipitates harvested from P removal by reusing FGD effluent at different pH values magnified at 100 k.

Effect of FPR on PRE

Effect of FPR on PRE and terminal pH values of the P-rich wastewater are shown in Figure 4. On one hand, the increase of FPR enhanced both Ca/P ratio and pH values, and favored P removal from P-rich wastewater. PRE was enhanced from 14.9% to 96.7% with FPR rising from 5% to 50%, and the terminal pH increased from 6.6 to 8.3. One-way factor analysis of variance showed that FPR had significant effects on both PRE and pH at p < 0.01 level. From Figure 4, FPR above 40% achieved PRE higher than 94%, and PO43−-P concentration in the effluent decreased to 2.0 mg/L. On the other hand, Ca/P ratio has a great effect on the structure of calcium phosphate precipitates. In the range of pH values, the major component of precipitates was TCP rather than HAP. Moreover, Mg2+ had a remarkable influence on the structure of precipitates.
Figure 4

Effect of FPR on PRE and terminal pH in P-rich wastewater.

Figure 4

Effect of FPR on PRE and terminal pH in P-rich wastewater.

Figure 5 presents the concentrations of Ca2+ and Mg2+ in raw (mixed) wastewater and supernatant after precipitation (precipitated effluent). With the increase of FPR, Ca2+ in the raw wastewater increased linearly owing to the rising proportion of FGD wastewater (Figure 5(a)). At FPR range of 5–20%, Ca2+ in the effluent increased at a lower slope, indicating that most Ca2+ was precipitated with P at low Ca/P ratio. The consumed Ca/P molar ratio rose from 2.20 at FPR of 5% to 4.89 at 20%, and then gradually decreased to 1.69 at 50%. The Ca/P molar ratio was higher than theoretical values of TCP (1.50) and HAP (1.67), indicating that Ca2+ was also consumed by other side reactions, such as the formation of CaSO4·0.15H2O. The tendency of Mg2+ is similar to Ca2+ (Figure 5(b)), revealing that Mg2+ was incorporated into calcium phosphate precipitates and promoted formation of TCP rather than HAP (Cao & Harris 2007).
Figure 5

Effect of FPR on concentrations of Ca2+ and Mg2+ in P-rich wastewater treated by reusing FGD effluent.

Figure 5

Effect of FPR on concentrations of Ca2+ and Mg2+ in P-rich wastewater treated by reusing FGD effluent.

Technical and economic analysis

Power plants are usually located at suburbs adjacent to wastewater treatment plants or industrial parks in many cities, and thus the FGD effluent can be piped to the plants or factories with P-rich wastewater (e.g. reject water from sludge treatment). Reusing FGD effluent as a calcium source not only saves the cost of precipitants potentially used for P removal, but also provides an economical substitute for current ZLD process by using membrane condensation and evaporation technology, which requires high energy consumption and capital costs.

Precipitants commonly used for chemical P removal were aluminum salt, with Al/P molar ratio of 2.5 (Ren et al. 2015). The unit price of polyaluminum chloride with Al2O3 content of 28% was $245 per ton, and cost of P precipitation (C1) was$0.68 per m3 for the P-rich wastewater in this study. For FGD wastewater to achieve current ZLD target, the operation cost consists of three parts: (1) reagents and energy consumption for pretreatment process to remove hardness ion (Ca2+, Mg2+ and sulfate), which is slightly higher than the current NPC process; (2) energy consumption for membrane condensation (C2), about $0.54 per m3 FGD wastewater for electrodialysis condensation; (3) steam and energy consumption for evaporation and crystallization process (C3), about$5.00 per m3 FGD wastewater. With the FPR of 40% to attain 95% PRE, the total cost saved by reusing FGD effluent for P removal is calculated as \$6.49 per m3 FGD wastewater according to Equation (1).
1

CONCLUSIONS

This study provided a technically and economically feasible and sustainable way for reusing FGD effluent for P removal. FGD effluent contained high concentrations of calcium and alkalinity, which obstructed its reclamation and reuse but was in favor of P precipitation. Increasing FPR achieved higher pH value and Ca/P ratio, and thus enhanced PRE to 94.3% at ratio of 40%. XRD and SEM analysis of harvested precipitates showed that increasing pH from 8 to 10 induced the conversion of HAP to TCP, and then to whitlockite. This process not only saves the cost of agents for P removal, but also provides an economical alternative for current ZLD technology that requires high energy consumption and capital costs. The results can be beneficial for FGD wastewater treatment and P removal under more stringent discharge limits by cooperation of different industries.

ACKNOWLEDGEMENTS

We thank Shanghai Rising-Star Program (16QA1401900) and National Natural Science Foundation of China (51408352) for financial support of this study.

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