Mechanisms of Phosphorus Removal by Recycled Crushed Concrete

Due to urbanisation, there are large amounts of waste concrete, particularly in rapidly industrialising countries. Currently, demolished concrete is mainly recycled as aggregate for reconstruction. This study has shown that larger sizes (2–5 mm) of recycled concrete aggregate (RCA) removed more than 90% of P from effluent when at pH 5. Analysis of the data, using equilibrium models, indicated a best fit with the Langmuir which predicated an adsorption capacity of 6.88 mg/g. Kinetic analysis indicated the equilibrium adsorption time was 12 h, with pseudo second-order as the best fit. The thermal dynamic tests showed that the adsorption was spontaneous and, together with the evidence from the sequential extraction and desorption experiments, indicated the initial mechanism was physical attraction to the surface followed by chemical reactions which prevented re-release. These results suggested that RCA could be used for both wastewater treatment and P recovery.


Introduction
Concrete is a major construction material which includes significant embedded resources, these include 50% of all the raw materials used and 40% of the energy. After manufacture, the other 50% of the total waste was generated by the concrete industry [1]. Recently, large volumes of demolition concrete waste have been produced due to limited reuse of older concrete buildings. For example in 2014, China produced 1.5 billion tonnes of construction waste; concrete was 34% of this total and only about 5% was reused [2]. This compares with above 50% in developed countries. New and alternative uses for recycled concrete are a priority in China where, traditionally, construction waste has been disposed of in landfills or deposited on river banks. This disposal used large areas of land as well as causing environmental nuisance and complaints.
There has been previous work on using recycled concrete aggregate (RCA) for water treatment and one of the most researched applications has been as a filter media due to its surface roughness and desirable chemical content (e.g., Mg, Ca, Fe, Al). Li et al. (2007) [3], for example, used fly ash aerated crushed concrete at laboratory scale to achieve 95.6% total nitrogen removal from sewage and landfill leachate. Guo et al. (2009) [4] also used RCA as a wastewater filter and showed it could remove 37% of COD and 55% of total P. The mechanism has been reported to be based on the calcium, aluminium, and iron content which can bind with phosphate [5]. Phosphorus (P) is needed for optimum crop production and mineral phosphorus reserves are thought to be limited with no substitutes. Research in both the U.S. and Europe has shown mineral P resources may be depleted in a few hundred years and recovery of P is urgent to maintain agricultural productivity [6]. Furthermore, removal of P from wastewater to meet environmental quality standards is now a significant cost in chemicals and power,  Before the experiments, all the glassware and plastic devices were soaked in 10% nitric acid solution for 1 day and then rinsed before each test with 250 mL deionised water. Sieved (2-5 mm) and manually hammered RCA was used for the experiments. RCA was also scanned by an infinite focus microscope (Alicona) (see Figure 1). An infinite focus microscope can quantitatively measure surface texture or deviations in a surface character. Figure 1 shows RCA has a crude surface, which indicates more adsorption sites. The specific surface area of RCA was also identified as 35 m 2 /m 3 . The RCA was examined by EDS (ED3000) and composed of 56% oxygen, 18% Si, 10% Carbon, 10% Ca, 2% Al, 2% K, and 1% Mg. Erhlmeyer flasks, with measured amounts of dried RCA media were used with 100 mL of a known P concentration as a solution of Potassium orthophosphate. The flasks were run in a thermostatically controlled orbital shaker for the defined periods (Table 1). By operating the shaker at various temperatures (Table 1), the data could be used to determine thermodynamic parameters. The pH of the solution was altered using a pH meter (HANNA HI9812-5) with either 0.1 M HCl or NaOH. The final solutions were filtered by a 0.45 µm membrane filter and dissolved concentrations of P were determined by Inductively Coupled Plasma analysis (Shimadzu ICP-9000) according to the international standard method [13]. All tests were conducted in triplicate. The amount adsorbed was calculated using the formula below: where q e is the adsorption capacity (mg/g), C o is the initial concentration of solution (mg/L), C e is the concentration after adsorption (mL/L), W is the weight of adsorbent (g) and V is the volume of solution (L). shaker at various temperatures (Table 1), the data could be used to determine thermodynamic parameters. The pH of the solution was altered using a pH meter (HANNA HI9812-5) with either 0.1 M HCl or NaOH. The final solutions were filtered by a 0.45 µm membrane filter and dissolved concentrations of P were determined by Inductively Coupled Plasma analysis (Shimadzu ICP-9000) according to the international standard method [13]. All tests were conducted in triplicate. The amount adsorbed was calculated using the formula below: where q is the adsorption capacity (mg/g), C is the initial concentration of solution (mg/L), C is the concentration after adsorption (mL/L), W is the weight of adsorbent (g) and V is the volume of solution (L).

Desorption of P
The P-treated RCAs from the previous experiments, with 2 g of RCA, were elutriated with 100 mL deionised water in Erhlmeyer flasks. Tests were conducted in triplicate for 24 h mixing and residual P concentration analysed as before.

Desorption of P
The P-treated RCAs from the previous experiments, with 2 g of RCA, were elutriated with 100 mL deionised water in Erhlmeyer flasks. Tests were conducted in triplicate for 24 h mixing and residual P concentration analysed as before.

Fractionation of Inorganic Phosphorus
Sequential extraction to analyse the various P-forms was used to identify the chemical reaction that occurred between the surface of the media and the P solution. Analysis of the P species was first reported by Dean (1938) [14] in soil complexes. Dean defined two types of P according to whether they were extracted by acid or alkali. Chang and Jackson (1957) [15] refined this by using four extraction solutions: loosely bound phosphate (LBP), aluminium phosphate (Al-P), iron phosphate (Fe-P), calcium phosphate (Ca-P), and occluded phosphate (O-P). Most procedures for P characterisation are based on the Chang and Jackson (1957) method with modification or improvements. Normally, Ca and Mg-P are extracted together by HCl [16][17][18]. The Ca-P amount was determined separately and the Mg-P by subtraction. The modification of the Chang and Jackson (1957) [15] and Hartikainen (1979) [19] method was used (see Table 2). After the standard adsorption test, 2 g RCA was used and the analysis was carried out in triplicate.

Effect of pH of Solution on Sorption
The best adsorption was observed at pH 5.0, when the adsorption capacity of P was 0.85 mg/g and removal reached 93.5% (Figure 2). The trend shows a decline in P adsorption with increasing pH to the lowest capacity at pH 8.0. Agyeia et al. [21] also pointed out acidic pH was best for P removal. There are two possible mechanisms, one has been attributed to the double layer effect whereby acidic H + was attracted to the concrete surface by Ca, Al, and Mg hydroxide content creating a secondary positive layer to bind the negative orthophosphate. The second possibility was that the phosphate ions could be converted to their acidic forms (e.g., H 2 PO − 4 and HPO 2− 4 ) binding to the positive surface. The amount of these phosphate ions generated would be proportional to the acidity of the solution. The acidic pH could also help the cement release more Ca 2+ ions into the solution to react with the hydrogen phosphate causing deprotonation and precipitation of Ca 3 (PO 4 ) 2 . The increase in P removal above pH 9 could be due to a similar mechanism but from the formation of OH enriched complexes precipitating calcium phosphate.

Effect of Dose of Sorbent on Phosphorus Sorption
In this experiment, the optimum pH 5 was used and the results in Figure 3 show the amount of P adsorbed was proportional to the amount of RCA added. At the lowest dose (1 g), P removal was 55.5% but 2 g removed 95% of P, similar to the maximum achieved. The normalised adsorption capacity reduced with increasing RCA from 0.99 to 0.17 mg/g, as added RCA was increased from 2 to 10 g. Similar trends have been also reported by other authors [22,23]. Increasing the dose of sorbent increased the total removal of P by increasing the availability of sorption sites. Specific adsorption capacity is a measure of the amount of P bonded by a unit weight of sorbent. The competition by ions for the sites caused a decrease in the specific uptake once all the sites were filled and adsorption capacity decreased with increments in sorbent dose. The simpler models assume a fixed number of adsorption sites according to the molecular structure and the number of positive ions responsible for the binding.  Figure 4 shows that the initial concentration of P in the solution influenced the equilibrium uptake of P achieved. The adsorption capacity increased linearly with the initial P concentration from

Effect of Dose of Sorbent on Phosphorus Sorption
In this experiment, the optimum pH 5 was used and the results in Figure 3 show the amount of P adsorbed was proportional to the amount of RCA added. At the lowest dose (1 g), P removal was 55.5% but 2 g removed 95% of P, similar to the maximum achieved. The normalised adsorption capacity reduced with increasing RCA from 0.99 to 0.17 mg/g, as added RCA was increased from 2 to 10 g. Similar trends have been also reported by other authors [22,23]. Increasing the dose of sorbent increased the total removal of P by increasing the availability of sorption sites. Specific adsorption capacity is a measure of the amount of P bonded by a unit weight of sorbent. The competition by ions for the sites caused a decrease in the specific uptake once all the sites were filled and adsorption capacity decreased with increments in sorbent dose. The simpler models assume a fixed number of adsorption sites according to the molecular structure and the number of positive ions responsible for the binding.

Effect of Dose of Sorbent on Phosphorus Sorption
In this experiment, the optimum pH 5 was used and the results in Figure 3 show the amount of P adsorbed was proportional to the amount of RCA added. At the lowest dose (1 g), P removal was 55.5% but 2 g removed 95% of P, similar to the maximum achieved. The normalised adsorption capacity reduced with increasing RCA from 0.99 to 0.17 mg/g, as added RCA was increased from 2 to 10 g. Similar trends have been also reported by other authors [22,23]. Increasing the dose of sorbent increased the total removal of P by increasing the availability of sorption sites. Specific adsorption capacity is a measure of the amount of P bonded by a unit weight of sorbent. The competition by ions for the sites caused a decrease in the specific uptake once all the sites were filled and adsorption capacity decreased with increments in sorbent dose. The simpler models assume a fixed number of adsorption sites according to the molecular structure and the number of positive ions responsible for the binding.  Figure 4 shows that the initial concentration of P in the solution influenced the equilibrium uptake of P achieved. The adsorption capacity increased linearly with the initial P concentration from  Figure 4 shows that the initial concentration of P in the solution influenced the equilibrium uptake of P achieved. The adsorption capacity increased linearly with the initial P concentration from 5 to 30 mg/L, but the proportion of P removed reached maximum at an initial concentration of 15 mg/L. This suggests the simple, single layer, fixed number of adsorption sites model may not be suitable at higher P concentrations. This is in line with previous equilibrium adsorption capacity experiments, which suggest that higher solute concentrations could encourage other mechanisms such as greater boundary concentrations which in turn lead to double layer adsorption and complex formation [23,24]. 5 to 30 mg/L, but the proportion of P removed reached maximum at an initial concentration of 15 mg/L. This suggests the simple, single layer, fixed number of adsorption sites model may not be suitable at higher P concentrations. This is in line with previous equilibrium adsorption capacity experiments, which suggest that higher solute concentrations could encourage other mechanisms such as greater boundary concentrations which in turn lead to double layer adsorption and complex formation [23,24].

Equilibrium Studies
The equilibrium equations are shown in Table 3, and were used to determine the mechanisms of adsorption. The Langmuir and Freundlich isotherms gave the best fit (presented in Table 4), the Langmuir showing a maximum adsorption capacity of 6.88 mg/g. The Langmuir is the simplest isotherm, based on monolayer sorption onto a surface with a finite number of identical sites, homogeneously distributed over the sorbent surface. Others have also reported that the Langmuir isotherm was the more robust when modelling complex aqueous mediums. The Frumkin and BET equations were introduced as extensions to the Langmuir isotherm, the Frumkin to take account of solute-solute interaction at a non-ideal surface [25]. The BET was introduced to model multi-layer adsorption but, except for the first layer, it also assumes equal energies of adsorption for each layer and no interactions between layers [26]. The linear correlation of both the Frumkin and BET are weaker than the Langmuir ( Table 4). The Freundlich equation included empirical constants as an acceptance that the adsorption sites were heterogeneous. The Temkin and the Dubinin-Radushkevich (D-R) adsorption equations are based on the thermodynamics of adsorption. The Tempkin is based on the principle that adsorption reduces the energy or heat in the adsorbed layer and the D-R on a Gaussian energy distribution as expected from a heterogeneous surface. The D-R model was reported to be suited to high and medium concentration solutions such as these P adsorption experiments [27]. The adsorption energy (E) derived from the D-R model can be used to predict adsorption mechanisms. Typical results are between 1-16 kJ/mol with E values lower than 8 kJ/mol indicating physical sorption [28]. In this study, E was calculated to be 2.24 kJ/mol, which would therefore be considered as physical adsorption.

Equilibrium Studies
The equilibrium equations are shown in Table 3, and were used to determine the mechanisms of adsorption. The Langmuir and Freundlich isotherms gave the best fit (presented in Table 4), the Langmuir showing a maximum adsorption capacity of 6.88 mg/g. The Langmuir is the simplest isotherm, based on monolayer sorption onto a surface with a finite number of identical sites, homogeneously distributed over the sorbent surface. Others have also reported that the Langmuir isotherm was the more robust when modelling complex aqueous mediums. The Frumkin and BET equations were introduced as extensions to the Langmuir isotherm, the Frumkin to take account of solute-solute interaction at a non-ideal surface [25]. The BET was introduced to model multi-layer adsorption but, except for the first layer, it also assumes equal energies of adsorption for each layer and no interactions between layers [26]. The linear correlation of both the Frumkin and BET are weaker than the Langmuir (Table 4). The Freundlich equation included empirical constants as an acceptance that the adsorption sites were heterogeneous. The Temkin and the Dubinin-Radushkevich (D-R) adsorption equations are based on the thermodynamics of adsorption. The Tempkin is based on the principle that adsorption reduces the energy or heat in the adsorbed layer and the D-R on a Gaussian energy distribution as expected from a heterogeneous surface. The D-R model was reported to be suited to high and medium concentration solutions such as these P adsorption experiments [27]. The adsorption energy (E) derived from the D-R model can be used to predict adsorption mechanisms. Typical results are between 1-16 kJ/mol with E values lower than 8 kJ/mol indicating physical sorption [28]. In this study, E was calculated to be 2.24 kJ/mol, which would therefore be considered as physical adsorption. Table 3. Equilibrium study models [25][26][27][29][30][31].

Isotherm Models Linear Expression Associated Equations Parameters
Freundlich logq e = logK f + 1 n logC e K and n are empirical constants  Previous filtration studies using cement or concrete have been summarised in Table 5. The majority of these studies used finer particles (>1 mm) to adsorb P because of the greater surface area and consequent adsorption capacity available from smaller sized particles. The aerated concrete had the greatest surface area and gave the largest adsorption capacity. It was reported as too fragile for water treatment however, and small media would also be vulnerable to clogging and wash out, giving high operating costs. The results of this study suggest larger sizes could be used as a compromise for wastewater treatment. The cement studies-by Zheng et al. [32] and Renman and Renman [33]-used a similar size as used here (Table 5) but both materials were unstable. Most of these studies have operated under acidic pH (<3) to achieve maximum P adsorption capacity but these acidic conditions would be unrealistic for most wastewaters. This study has demonstrated that concrete demonstrated reliable P removal at a higher pH (pH 5) which could be achieved in the microenvironment of a wastewater filter.

Kinetic Analysis
Four kinetic models were also used to analyse the data ( Table 6). These were the standard first and second-order equations plus the Fractional power equation which is a modified form of Freundlich. Finally the Elovich equation was used since it had previously been used to study NH + 4 removal by Chien and Clayton (1980) [38]. The results and correlation coefficients (R 2 ) are summarised in Table 7. The best fit was pseudo second-order followed by the first order equation. The experimentally measured adsorption capacity (0.749 mg/g) was between those calculated from the pseudo first-and second-order equations. The experimental results indicated that RCA adsorbed most when at high P concentrations; the calculated maximum capacity suggested adsorption could reach 0.9 mg/g. Table 6. Kinetic study models [39,40].

Kinetic Study Linear Expression Parameters
The pseudo-first-order log(Q e − Q t ) = −   Figure 5 shows the adsorption rate and removal efficiency increased until equilibrium was reached at 93% P adsorbed after 12 h and 99% at 24 h. The results were carried out at typical domestic wastewater P concentrations (around 10-15 mg/L) and demonstrated a rate of adsorption likely to be achievable in a wastewater treatment plant without excessive costs.  Figure 5 shows the adsorption rate and removal efficiency increased until equilibrium was reached at 93% P adsorbed after 12 h and 99% at 24 h. The results were carried out at typical domestic wastewater P concentrations (around 10-15 mg/L) and demonstrated a rate of adsorption likely to be achievable in a wastewater treatment plant without excessive costs.

Effect of Temperature on Sorption
The influence of temperature on the thermodynamics of adsorption (Gibbs free energy (∆G • ), Enthalpy change (∆H • ), and Entropy change (∆S • )) was calculated according to the procedure described by Tosun (2012) [28] as shown in Equations (1)- (3): where R = gas constant (8.314 J/mol K); T = absolute temperature; b = the distribution coefficient, which was calculated by: where Ca = equilibrium concentration of P on adsorbent (mg/L). The relation between (∆H • ), (∆G • ) and (∆S • ) was given by: (3) Figure 6 shows that the adsorption of P increased with rises in temperature and that the adsorption process was endothermic. It is suggested that the higher temperature resulted in swelling of the RCA, creating more space for adsorption as well as accelerating molecular diffusion and the chemical reaction. The b value was obtained from ln( Q e C e ) vs. Q e and by extrapolating Q e to zero to obtain the intercept value. Results for ∆G • , ∆H • , and ∆S • were calculated from the slope and intercept of a plot of ln K vs. (1/T) (see Figure 7).
The thermodynamic parameters are summarised in Table 8. The negative values of ∆G • indicate spontaneous adsorption of P on the RCA and the decreasing values of ∆G • confirm the experimental data showing acceleration at higher temperatures. The ∆H • is positive and less than 20 kJ/mol, while the value of ∆H • is less than 20 kJ/mol, indicating the adsorption process is physical [41]. It has been suggested that the positive value of ∆S • is due to the release of the water from phosphate as it is adsorbed onto the surface to increase the degrees of freedom and entropy of the system [42,43]. Other thermodynamic sorption studies for P are summarised in Table 9. The free energy range of RCA was similar to dolomite and activated carbon. Adsorption to these natural materials were also spontaneous, endothermic, and with increased entropy. The similarity to activated carbon also suggests non-specific adsorption sites corroborating the results from the isotherm and kinetic studies. adsorption process was endothermic. It is suggested that the higher temperature resulted in swelling of the RCA, creating more space for adsorption as well as accelerating molecular diffusion and the chemical reaction. The b value was obtained from ln( ) vs. Q and by extrapolating Qe to zero to obtain the intercept value. Results for ∆G°, ∆H°, and ∆S° were calculated from the slope and intercept of a plot of ln K vs. (1/T) (see Figure 7).  The thermodynamic parameters are summarised in Table 8. The negative values of ∆G° indicate spontaneous adsorption of P on the RCA and the decreasing values of ∆G° confirm the experimental data showing acceleration at higher temperatures. The ∆H° is positive and less than 20 kJ/mol, while the value of ∆H° is less than 20 kJ/mol, indicating the adsorption process is physical [41]. It has been suggested that the positive value of ∆S° is due to the release of the water from phosphate as it is adsorbed onto the surface to increase the degrees of freedom and entropy of the system [42,43]. Other thermodynamic sorption studies for P are summarised in Table 9. The free energy range of RCA was similar to dolomite and activated carbon. Adsorption to these natural materials were also spontaneous, endothermic, and with increased entropy. The similarity to activated carbon also suggests non-specific adsorption sites corroborating the results from the isotherm and kinetic studies.    3.2. Desorption of Phosphorus Figure 8 shows the proportion of P re-released. The desorption was in the range of 4-7% and lowest at the higher initial P concentrations with maximum desorption at the initial P = 10 mg/L. The adsorption and desorption capacities were not equal, corroborating a previous hypothesis that different mechanisms were involved at different P concentrations and that some were irreversible. It is suggested that at the lower concentrations adsorption was due to just physical adsorption, which then enabled easier P release. Higher concentrations of P, on the other hand, could be sufficient to cause reactions with the metals such as complexation and precipitation; desorption would therefore be reduced [47]. A lower desorption rate at smaller P concentrations, was also reported by Yu et al. (2009) [35]. They tested desorption from mortar (cement mixed with sand) and found the release was 9% but they deduced that the adsorption was mainly due to chemical reaction. is suggested that at the lower concentrations adsorption was due to just physical adsorption, which then enabled easier P release. Higher concentrations of P, on the other hand, could be sufficient to cause reactions with the metals such as complexation and precipitation; desorption would therefore be reduced [47]. A lower desorption rate at smaller P concentrations, was also reported by Yu et al. (2009) [35]. They tested desorption from mortar (cement mixed with sand) and found the release was 9% but they deduced that the adsorption was mainly due to chemical reaction.

Fractionation of Inorganic Phosphorus
Concrete is composed of Ca, Mg, Si, and Al hydroxides [33,37,48]. The proportions of each metal bound P types were measured. The Si was assumed to be inactive and not analysed. The data is shown in Figure 9: in the raw concrete O-P was the predominant form of P (37%), the remainder was mostly Ca-P (29%) combined with the Al-P, Mg-P and labile P were less than 10%. There was no Fe-P observed in this study. In the used RCA, on the other hand, labile P (43.3%) was the highest proportion, O-P remained high (25%), Al-P and Ca-P were reduced but Mg-P increased, possibly reflecting variations in the mineral composition. Previous work indicates P reacts with calcium and magnesium, from the cement, to form insoluble precipitate (detail see Table 10).

Fractionation of Inorganic Phosphorus
Concrete is composed of Ca, Mg, Si, and Al hydroxides [33,37,48]. The proportions of each metal bound P types were measured. The Si was assumed to be inactive and not analysed. The data is shown in Figure 9: in the raw concrete O-P was the predominant form of P (37%), the remainder was mostly Ca-P (29%) combined with the Al-P, Mg-P and labile P were less than 10%. There was no Fe-P observed in this study. In the used RCA, on the other hand, labile P (43.3%) was the highest proportion, O-P remained high (25%), Al-P and Ca-P were reduced but Mg-P increased, possibly reflecting variations in the mineral composition. Previous work indicates P reacts with calcium and magnesium, from the cement, to form insoluble precipitate (detail see Table 10). shown in Figure 9: in the raw concrete O-P was the predominant form of P (37%), the remainder was mostly Ca-P (29%) combined with the Al-P, Mg-P and labile P were less than 10%. There was no Fe-P observed in this study. In the used RCA, on the other hand, labile P (43.3%) was the highest proportion, O-P remained high (25%), Al-P and Ca-P were reduced but Mg-P increased, possibly reflecting variations in the mineral composition. Previous work indicates P reacts with calcium and magnesium, from the cement, to form insoluble precipitate (detail see Table 10). Figure 9. The comparison of different type of P before and after test. Figure 9. The comparison of different type of P before and after test.

Sequence
Chemical Formula In this study, Mg-P was low in the raw concrete and there are few studies of Mg-P due to its high water solubility. This solubility and reactivity of the Mg-ion compared to Ca, Fe, and Al means Mg-P forms rapidly, but is less crystalline and vulnerable to remobilisation by a wide range of conditions (including pH). Ca-P is more stable due to slower crystallisation [49]. Previous work has also indicated the form of P does vary according to the chemical composition of the filter media (Table 11). Oyster shell, for example, was more than 50% CaO which then formed the majority of the adsorbed P as Ca-P [50]. Zeolite was reported to be 70% SiO 2 , 10% Al 2 O 3 , 2.5% CaO, and 1.5% Fe 2 O 3 [51][52][53]. The inactivity of Si would explain its poor performance for P removal. In other materials-such as drinking water treatment residues, fly ash, and Bauxite residues-the total P amount was higher than RCA, but the soluble P and sizes available were much smaller, and were not convenient as wastewater filter media.
The used RCA contained 43% soluble P which is the second highest when compared to previous work and blast furnace slag contained most soluble P (see Table 11). Hylander (1999) [58] carried out loading scale pot experiments and stated that slag sorbed P was highly plant available. Therefore, RCA could be used as fertiliser. This is due to the fact that soluble P could be recovered through equilibrium soluble P-concentration, pH change, or redox potential [59,60]. It was confirmed in previous tests that soluble P could be re-released as it was physical attraction, corroborating the predictions of the isotherm, kinetic, and desorption tests. The reaction time was important because previous work has shown that the surface-attracted P might gradually become more permanently bound through complexation and precipitation reactions [54,61]. This was demonstrated in Table 11 which showed adsorption was followed by the formation of apatite complexes. O-P was the second highest indicating that between a third and a quarter of the P became fixed immediately, and would resist re-release, but it still could be re-converted into available P by soil microorganisms (e.g., Phosphate solubilising bacteria) [62].

Conclusions
This study tested recycled concrete for P removal from wastewater, using a larger sized (2-5 mm) and therefore more suitable media than previous studies in practice. More than 90% P removal was achievable using concentrations typical of sewage effluents (p < 15 mg/L) with an equilibrium time of 24 h. Model analysis predicted that capacity could reach 6.88 mg/g. The experiments using an analysis of the kinetics (pseudo second-order R = 0.9916) and thermodynamics indicated the adsorption process was spontaneous and, together with the differential extraction and desorption tests results, indicated the mechanism was surface physical electrostatic attraction followed by chemical precipitation. The results suggested that half the P adsorbed to the surface was labile and available for recovery, or, after further crushing, could be used directly as a fertilizer supplement. Further studies on the uptake and impact on plant growth from P enriched RCA would be useful.
The results offer a new usage for high cement RCA which is undesirable for aggregate, due to a lower density, high water adsorption, and reduced structural performance [63,64]. Further work is also suggested to measure how concrete adsorbs heavy metals and persistent organics and how these might interact with C, P, and N in effluents.