Synergistic Removal of Zinc and Copper in Greenhouse Waste

Dec 30, 2015 - Aerated Fluidized Bed Treatment for Phosphate Recovery from Dairy and Swine Wastewater. Alon Rabinovich , Ashaki A. Rouff , Beni Lew , ...
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Research Article pubs.acs.org/journal/ascecg

Synergistic Removal of Zinc and Copper in Greenhouse Waste Effluent by Struvite Ashaki A. Rouff,* Marlon V. Ramlogan, and Alon Rabinovich Department of Earth and Environmental Sciences, Rutgers University, 101 Warren Street, Newark, New Jersey 07102, United States S Supporting Information *

ABSTRACT: Struvite (MgNH4PO4·6H2O) can be recovered from wastewaters for mitigation of phosphorus content. However, the interaction of dissolved constituents with struvite is rarely evaluated. Removal of heavy metals and total organic carbon (TOC) in a greenhouse wastewater (GW) by struvite was investigated. Presynthesized struvite was added to GW and removal of Zn (689 μg/L), Cu (151 μg/L), and TOC (51 mg/ L) monitored from 1 to 26 d. Metal uptake in sodium nitrate solutions was used to assess competition, and the influence of other GW constituents on sorption. Struvite was also directly precipitated from GW (PPT). Recovered GW solids had 64− 247 mg/kg Zn, 12−54 mg/kg Cu, and 1721−8806 mg/kg TOC, with lowest loadings for PPT and highest for 26 d solids. X-ray absorption spectroscopy detected polymerized Znphosphate, induced by dissolved phosphorus in GW, and Cu copolymerization, initially limited by aqueous Cu-organic complexation. Sorbed Cu shifts Fourier transform infrared-sensitive phosphate bands and changes the intensities of reflections in X-ray diffraction patterns of struvite more so than Zn. Struvite from GW is more susceptible to thermal decomposition than unreacted struvite, evolving CO(g), CO2(g), NH3(g), and H2O(g). Therefore, struvite from GW sorbs metals and organics, and can release sorbed and structural components to the aqueous and gas phases. KEYWORDS: STA-FTIR, XAFS, FTIR, Organic carbon, Greenhouse gases, Struvite, Heavy metals

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INTRODUCTION Struvite (MgNH4PO4·6H2O, MAP) can be recovered from nutrient-rich sources such as wastes and wastewaters to mitigate P content.1 The mineral product can subsequently be used as a sustainable source of phosphate fertilizer;1−3 or disposed of if the goal is solely to reduce wastewater contamination by P. The recovery process and optimization thereof from both simulated4,5 and actual wastes6−11 has been assessed. However, the association of additional constituents in wastes with struvite is less well documented. Though the interaction of contaminants such as heavy metals has been acknowledged, and to some extent quantified,8,12 there are few fundamental studies evaluating the mechanisms of sorption.13−16 This is critical for determination of subsequent mobility of contaminants, and the fate of the recovered P. Furthermore, the impact of sorption on the mineral structure and properties remain relatively unstudied16 but can provide insight into the mineral behavior when repurposed or discarded. Optimized conditions for struvite crystallization enhance contaminant sorption, which even at low concentrations affect the mineral structure. The potentially toxic elements As and Cr were observed to incorporate into the mineral structure13 or form surface-bound precipitates14 with the sorption mechanism predictive of subsequent contaminant release. The micronutrient Zn sorbed by variable mechanisms, contingent upon © XXXX American Chemical Society

initial concentration and presence after or during struvite mineralization.16 For adsorption, Zn formed tetrahedral complexes that polymerized to a precipitate phase, whereas in coprecipitation-type experiments Zn was in octahedral coordination, and the precipitate formed at high concentration retained structural interactions with struvite. Fourier transform infrared (FTIR) spectroscopy revealed Zn significantly distorted struvite phosphate tetrahedra at low concentrations in adsorption-type experiments.16 Contingent upon initial concentration and precipitation conditions, As, Cr, and Zn sorbed to struvite at concentrations in excess of recommended fertilizer limits.13,14,16 Struvite can undergo solid-state decomposition, releasing structural ammonium and water to the gas phase, at both ambient and elevated temperature.17−19 Though thermal analysis requires application of temperatures >25 °C, response to thermal stress is a good indicator of overall and relative stability, and is applicable to ambient processes, which primarily occur at slower rates. Thermal decomposition of struvite, including pathways and gaseous and solid products, has been assessed,19,20 and the impact of pH of crystallization on Received: October 21, 2015 Revised: December 16, 2015

A

DOI: 10.1021/acssuschemeng.5b01348 ACS Sustainable Chem. Eng. XXXX, XXX, XXX−XXX

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ACS Sustainable Chemistry & Engineering susceptibility to decomposition delineated.5 However, any correlation between contaminant content and the decomposition process remains relatively unexplored.21 Sorbates that affect struvite decomposition may also influence the propensity for volatile release, thereby introducing components to the atmosphere in addition to soil−water systems. This is environmentally relevant as H2O(g) is a greenhouse gas, and NH3(g) is a pollutant associated with N-bearing fertilizers and animal wastes.22−24 Furthermore, large quantities of struvite can be recovered from wastewaters. For example, >90% of 1400 mg/L ammonium in an industrial wastewater effluent was recovered as struvite;25 and a wastewater treatment plant pilot study projected recovery of ∼1000−1200 t of struvite annually from digestor supernatant.26 Therefore, when scaled-up, the volatile content of recovered struvite could be substantial, and the potential for gas release requires assessment. Greenhouse waste has been evaluated for P recovery as struvite27 as discharged effluent can cause pollution and eutrophication of receiving waters. Effluent from open hydroponic greenhouses can have high concentrations of P and N,28−30 and recirculation and recycling of water in closed systems can lead to salt and pathogen buildup in hydroponic solutions, ultimately requiring discharge.29 Nutrient loads in greenhouse effluent can be reduced by constructed wetlands31,32 and denitrification filters,33 however precipitation of struvite can also be used to recover nutrients,27 thereby maximizing the environmental benefit. In the current study, the interaction of struvite with dissolved constituents in a greenhouse waste was explored. Though several dissolved metals were present, Zn and Cu, both micronutrient metals, were detected at highest concentrations. Therefore, the simultaneous removal of these metals was monitored. Experiments using simplified electrolyte-based solutions at the same ionic strength and metal concentration were conducted in parallel. Because it is difficult to characterize completely all constituents in a waste source, these solutions were used to delineate the impact of organics and unidentified constituents in the waste on metal uptake. Experiments at variable Zn:Cu ratios were used to delineate competitive or synergistic effects during sorption. The extent and mechanisms of metal sorption, the influence of organics, and the effect on the mineral functional groups were evaluated. The thermal stability relative to unreacted struvite, and susceptibility to release of gaseous decomposition products was also assessed. The research provides additional insight into contaminant interaction and release by struvite that has been in contact with waste sources.

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measured at 50 mg/L, with a SUVA of 3.5 L/(mg·m) in the range of fulvic acids.34 The collected solution was filtered using a 0.45 μm filter, and a mass of precharacterized struvite suspended at a loading of 1 g/L and reacted for 7 d. A batch of unfiltered solution was also reacted with 1 g/L struvite for 7 d. All experiments were prepared in duplicate. After addition of struvite, the pH rose to 7 and remained constant. Aliquots of solution were sampled intermittently over the 7 d reaction for analysis of total aqueous Zn and Cu concentration by ICP-OES. Both filtered and unfiltered GW solutions without struvite were sampled as blanks. A second set of experiments was prepared for collection of solids for subsequent analysis. Struvite was suspended in 0.45 μm filtered GW solution at a loading of 1 g/L and reacted for 1, 7, or 26 d. Longer reaction times were implemented in part to increase contaminant loadings and therefore investigate this effect. An additional experiment was conducted to precipitate struvite directly from the GW solution (PPT) to assess metal interaction under these conditions (Supporting Information). The solids were recovered by filtering the entire solution volume using a 0.45 μm filter, and allowed to air-dry. A fraction of the solid was acid-digested, and the metal concentration, TOC concentration, and SUVA were determined as described above. The remaining solid was retained for further characterization. Sodium Nitrate Solutions. On the basis of the composition of GW, simplified sodium nitrate (NN) solutions were made at pH 7 by addition of 0.1 M NaNO3 to deionized water (DI). Stock solutions of ZnSO4 and CuCl2 were used to add 10 μM Zn and 2 μM Cu, so initial metal concentrations were consistent with the GW solution. Struvite was added at a loading of 1 g/L and reacted for 1, 7, or 26 d. Additional NN solutions were prepared to assess the behavior of Zn and Cu at different ratios, and in single metal systems, in part to discern competitive/synergistic effects. The initial Zn:Cu concentration ratios were 10:10, 10:0, 0:2, 0:10 μM, and samples were reacted with struvite for 1 d. Aqueous samples and reacted solids were collected and treated as described above for GW solutions. ATR-FTIR Analysis. FTIR spectroscopy data were collected for recovered solids using a PerkinElmer Spectrum 100 instrument and a universal attenuated total reflectance (ATR) accessory with a ZnSe crystal. Spectra were collected in the 600−4000 cm−1 range, with a resolution of 4 cm−1. For each sample, 20 scans were collected and averaged to produce the final spectrum. XAFS. X-ray absorption fine structure spectroscopy (XAFS) data for select solids were collected at beamline 12BM-B, Advanced Photon Source (APS), Argonne National Laboratory (ANL), Argonne IL. The beamline was equipped with a Si(111) monochromator tuned to the Zn K-edge (9659 eV) and the Cu K-edge (8979 eV). Data were collected in fluorescence mode using a 13-element Ge detector. To reduce the elastic scattering, a Cu filter for Zn, and a Ni filter for Cu measurements was placed between the sample and detector. Data reduction and analysis was performed using Athena and Artemis software included with the IFEFFIT program.35,36 STA-FTIR. Solids were characterized by simultaneous thermal analysis (STA) coupled with FTIR for evolved gas analysis (EGA). A Netzsch Perseus instrument consisting of a STA 449 coupled to a Bruker Alpha FTIR was used, with N2(g) as both protective and sample purge gas. Solids were heated from 25 to 500 °C at a rate of 10 °C/min and infrared spectra of evolved gases were collected every 12 s over a range of 600−4000 cm−1 at 8 cm−1 resolution.

MATERIALS AND METHODS

Greenhouse Waste Solutions. Greenhouse waste (GW) was collected from a greenhouse on Long Island, New York. The waste was runoff from plants watered with N-P-K (20-10-20) fertilizer. The initial pH was 6. Trace metal and major cation concentrations were measured by inductively coupled plasma optical emission spectroscopy (ICP-OES) using a PerkinElmer Optima DV-5300 instrument. The major anion concentration was measured by Ion Chromatography (IC) using a Dionex DX 500 instrument with an IonPac As14 universal ion-exchange column. Total organic carbon (TOC) was determined colorimetrically using a Hach DR 3900 spectrophotometer, and correlated with the absorbance at 254 nm using an Agilent Cary 300 spectrophotometer for specific ultraviolet absorbance (SUVA) (Supporting Information). The concentrations of the dominant trace metals were 10.5 μM Zn and 2.4 μM Cu, and the overall ionic strength was calculated to be ∼0.1 M with NO3−(aq) as the dominant anion (Table S1). The TOC concentration was

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RESULTS AND DISCUSSION Metal Uptake and Impact of GW Constituents. The metal concentrations remaining in GW solutions, as a percentage of dissolved metal in corresponding blank (struvite-free) solutions, is reported in Figure 1. Removal of metals increases over time with 49−53% Zn and 84−90% Cu remaining in solution at 7 d. The higher fraction of Zn removal may be due to differences in speciation of the two metals. The metal speciation was determined using Visual Minteq (Table B

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increase from 111 to 247 mg/kg Zn and from 30 to 54 mg/kg Cu from 1 to 26 d; and approach the initial Zn/Cu ratio in solution of 4.6 for GW 7 d and 26 d solids, indicating that uptake is not competitive. Organic matter is also associated with the solids, but cosorption with metals cannot fully account for the observed loadings, thus sorption may occur independently of metals. Precipitated struvite, GW PPT, has lower metal loadings at 64 and 12 mg/kg Zn and Cu, but comparable TOC content to the GW 1 d solid. Metal loadings for solids from NN solutions also increase with time (Table 1), and exceed those from GW solutions by a factor of 1.5−2 for both metals. This may be due to the absence of organics and other complexing ligands, and therefore dominance of the free metal, 91% Zn2+(aq) and 78% Cu2+(aq) (Table S2), available for sorption. The metal loadings for NN 1 d solids with variable solution Zn:Cu concentrations are reported in Table 1. When Zn alone is present, the loading at 143 mg/kg, is slightly higher than 127 mg/kg for the 1 d Zn:Cu 10:2 solid (NN 1 d, Table 1a). However, addition of an equimolar concentration of Cu in the Zn:Cu 10:10 solution increases the Zn loading to 377 mg/kg, with a Cu loading of 385 mg/kg. This alludes to a synergistic cosorption process, in that the presence of Cu enhances Zn uptake. For the 10:2 solution, Zn cosorption may actually be limited by low Cu when compared to sorption of Zn alone in the Cu-free solution. The Cu loading for the 0:10 solid, 384 mg/kg, is the same as for the 10:10 solid, so Cu sorption is unaffected by Zn. Loadings are also similar for 10:2 and 0:2 solids at 46 and 52 mg/kg Cu, indicating no significant impact of Zn even at lower Cu concentration. Metal Binding Configuration. XAFS was used to determine the metal binding configuration in the solids (Table 2, Table S3−S5, Figure S1−S2). For all Zn-bearing solids, the first-shell Zn−O distance is 1.95 Å indicating tetrahedral coordination. Solids from GW solutions have similar structure regardless of reaction time with P, Zn, and O shells at 3.06−3.07, 3.28−3.32, and 3.71−3.74 Å. The structure is consistent with a mixed-phase hydrated Zn-phosphate similar to hopeite (Zn3(PO4)2·4H2O), and liversidgeite (Zn6(PO4)4· 7H2O).16 Formation of this phase may be accelerated by the presence of ∼1 mM P-PO437 in the GW solution. For the 26 d

Figure 1. Percent Zn or Cu remaining in solution over a 7 d reaction period for 0.45 μm filtered and unfiltered GW solutions in contact with 1 g/L struvite.

S2). The dominant Zn species is 61% Zn2+(aq); however, ∼70% Cu is complexed with organic matter, with only 11% Cu2+(aq) available (Table S2). The calculated percent Zn2+(aq) and Cu2+(aq) species corresponds with the percent removal from solution, suggesting preferred interaction of the free metal with the solid. This correlation also suggests that organic parameters included in the Visual Minteq database are sufficient to describe metal−organic complexation in the GW solution. In unfiltered GW solutions, metal removal is slightly enhanced, which may be due to recovery of metal associated with particles >0.45 μm with the solid. Overall, the impact of filtration on metal removal is minimal, thus primarily dissolved metal species interact with the solid. As the concentration of Zn and Cu removed from solution increases with time, so do the metal loadings on the solids recovered from filtered GW solutions (Table 1). The loadings

Table 1. Initial Aqueous Concentrations and Final Solid Loadings of Zn and Cu time dependent GW and NN experimentsa GW solutions aqueous solid 1d 7d 26 d PPT

Zn (ug/L, μM) 689, 10.5 Zn (mg/kg) 111.3 155.3 246.8 63.9

Cu (ug/L, μM) 151, 2.4 Cu (mg/kg) 29.8 32.3 53.7 11.8

NN solutions

Zn/Cu TOC (mg/L) 4.6 51 Zn/Cu TOC (mg/kg) 3.7 1825 4.8 4041 4.6 8806 5.4 1721 additional 1 d NN experimentsb

aqueous Zn:Cu 10:10 10:0 0:2 0:10

Zn (μM) 10 10 0 0

Cu (μM) 10 0 2 10

Zn (μM) 10 Zn (mg/kg) 127.4 261.9 431.4

Cu (μM) 2 Cu (mg/kg) 45.9 69.7 99.2

Zn/Cu 5 Zn/Cu 2.7 3.7 4.2

solid Zn/Cu 1.0

Zn (mg/kg) 377.0 142.9

Cu (mg/kg) 384.9

Zn/Cu 1.0

51.9 384.4

a

After 1, 7, or 26 d reaction time for GW and NN samples, with organic concentrations included for GW samples. bNN samples reacted for 1 d at variable Zn:Cu ratios. C

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ACS Sustainable Chemistry & Engineering Table 2. Results from Fits to Zn and Cu EXAFS Data for Solids Recovereda time dependent GW and NN experimentsa GW solutions shell Zn: 1 d

Zn: 7 d

Zn: 26 d

a

CN

b

0.1 M NN solutions

R (Å)

σ (Å )

shell O Zn P

4 1 1

1.95 3.25 3.58

0.005 0.007 0.004

O Zn Cu O O Zn Cu O O P Cu O P Cu O P Zn O

4 1 1 1 4 1 1 1 4 2 2 4 2 1 4 1 1 4

1.95 3.25 3.41 3.74 1.95 3.24 3.40 3.75 1.96 3.28 3.35 1.96 3.28 3.33 1.96 3.16 3.48 3.68

0.006 0.008 0.010 0.010 0.006 0.008 0.009 0.010 0.003 0.019 0.013 0.004 0.011 0.006 0.004 0.009 0.010 0.010

c 2

2

O P Zn O O P Zn O O P Zn O

4 1 1 2 4 1 1 2 4 1 1 2

1.95 3.06 3.30 3.74 1.95 3.09 3.28 3.74 1.95 3.07 3.32 3.72

0.004 0.011 0.009 0.006 0.004 0.007 0.010 0.006 0.005 0.006 0.010 0.009

O

4

1.92

0.003

Cu: 1 d

Cu: 7 d

Cu: 26 d

a

CN

b

R (Å)

σ (Å2)

c 2

additional 1 d NN experimentsb Zn EXAFS a

Zn:Cu

shell

10:0 or 0:10

O Zn O

4 1 1

10:10

O Zn Cu O

4 1 1 1

CN

Cu EXAFS σ (Å2)

shell

1.95 3.27 3.76

0.005 0.008 0.007

1.95 3.23 3.39 3.75

0.007 0.008 0.007 0.015

O P Cu O O P Cu O O P Cu

b

R (Å)

c 2

0:2

a

a

CN

b

4 2 1 2 4 2 1 2 4 1 2

R (Å) 1.96 3.20 3.25 3.79 1.95 3.19 3.23 3.77 1.96 3.29 3.36

σ (Å2)

c 2

0.004 0.007 0.005 0.004 0.005 0.008 0.010 0.006 0.004 0.003 0.008

From GW and NN solutions reacted for 1, 7, or 26 d. bFrom NN solutions reacted for 1 d at variable Zn:Cu ratios.

sample, the last two shells can also be fit with Zn at 3.23 Å and Cu at 3.39 Å (Table S3), indicating some cosorption of Cu with the samples and that Zn may be in more than one coordination environment. Therefore, at longer reaction times sufficient Cu sorption may occur for detection of this shell in the GW samples by XAFS. When precipitated from solution, as for the GW PPT solid, Zn has a slightly relaxed first-shell O bond length of 1.98 Å, and is polymerized, with a Zn shell at 3.26 Å (Table S3). For NN samples, a Zn shell is detected at 3.24−3.25 Å. At 1 d, a P shell at 3.58 Å indicates monodentate complexation of Zn polymers with struvite phosphate tetrahedra, and therefore direct linkages with the mineral structure.16 At 7−26 d, a Cu shell is present at 3.40−3.41 Å due to cosorption of the metals. The overall structure for these samples is similar to the 10:0 NN solid, except for the presence of Cu, and when Cu is present at 10:10 the structures are identical. Therefore, Zn may initially interact directly with struvite due to faster sorption (Figure 1), and higher initial concentration compared to Cu.

Then as Cu sorption increases with time, Zn becomes increasingly coprecipitated with Cu. The Zn−Cu distance is similar to that in veszelyite (Cu,Zn)2ZnPO4(OH)3·2H2O (3.44 Å),38 and the Zn−Zn distance in hopeite (3.39 Å).39 Because the Cu concentration is likely too low for veszelyite to be a dominant phase, Cu may be substituting for Zn in a hopeitelike structure. The 7 d sample second and third shells can also be fit with P at 3.11 Å and Zn at 3.28 Å (Table S3), which is similar to the equivalent sample from the GW solution, and consistent with a mixed Zn-phosphate configuration. Because there is no initial P in the NN solutions, the formation of Znphosphate may be limited until sufficient dissolved P from the mineral has accumulated so that the phase is detectable in the 7 d NN sample; after which, cosorption with Cu dominates. Comparing all available fits for the GW and NN 7 d and 26 d samples (Table S3), and noting the presence of Zn in more than one coordination environment, there are consistencies in the structural data for samples reacted at equivalent times. D

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Figure 2. FTIR spectra for unreacted struvite (MAP) and struvite solids in contact with GW and NN solutions for (a) 1, 7, or 26 d with dashed lines indicating the positions of the ν3 PO43− (1000 cm−1) and ν4 NH4+ (1450 cm−1) bands; (b) 1 d of reaction with Zn:Cu 10:0, 0:2 in the region of the phosphate band.

at 970 and 1020 cm−1, and a small shoulder at 1060 cm−1.16 The band centered at 1450 cm−1 is attributed to NH4+ ν4 asymmetric bending, and at ∼760 cm−1 to H-bonding in water.43 For GW 1−7 d solids, the peak at 1020 cm−1 for the ν3 PO43− band increases in intensity and is resolved into a single peak centered at 1000 cm−1. The shoulder at 1060 cm−1 also increases in intensity. At 26 d of reaction, this band broadens significantly with additional growth of the shoulder at 1060 cm−1 and an overall shift to higher wavenumbers; and the ν4 NH4+ band and bands in the 800−650 cm−1 range disappear. This is different for the NN solutions, which show growth of the peak at 1020 cm−1 after 1 d of reaction, with no discernible change over time, and no impact on the ν4 NH4+ band. The disappearance of the ν4 NH4+ band for the GW 26 d solid is not due to a change in mineralogy,20 as evidenced by the XRD pattern (Figure S3). Therefore, this feature along with broadening of the phosphate band may be due to the association of organics. The 1 d NN solutions provide some insight into the impact of Zn and Cu on the ν3 PO43− band (Figure 2b). Comparing 10:0 and 0:2 solids to unreacted struvite, both Zn and Cu shift the band to higher wavenumbers with growth of the peak at 1020 cm−1 and reduced intensity of the peak at 880 cm−1. Although these changes are minimal for the 10:0 Zn sample, the impact of Cu, even though the concentration is lower, is more pronounced. This may be due to higher affinity of Cu for phosphate based on electronegativity, hydrolysis and stability constants of Cu and Zn with this ligand.42 That Cu preferentially interacts with the struvite structure is also evident from XAFS that reveals bidentate linkages to struvite phosphate tetrahedra. Additionally, changes in the intensity of reflections in the XRD patterns of Cu-sorbed solids at 14−16 and 20−22° 2θ, suggests a structural effect of Cu on the planes associated with these reflections in struvite (Figure S3). The shape of the ν3 PO43− band for the 10:2 1 d NN solid can be described by both the 10:0 and 0:2 solids. Comparing this to the GW 1 d solid, the features in this region of the FTIR spectrum can also be attributed to the interaction of Zn and Cu. However, the peak at 760 cm−1 decreases in intensity along

Cu XAFS for GW solids was only possible for the 26 d sample as other loadings were below detection. Because of the quality of the data, only a first-shell fit was possible generating a Cu−O distance of 1.92 Å for tetrahedral coordination. For NN samples, tetrahedral complexes with a 1.96 Å distance were detected. Though Jahn−Teller distorted octahedrally coordinated Cu complexes is common for sorption on carbonates,40 tetrahedral complexes have been observed on oxide surfaces.41,42 At 1−7 d, P at 3.28 Å and Cu at 3.33−3.35 Å indicates complexation of Cu polymers with phosphate tetrahedra. The structural data is the same as the 0:2 NN sample, so the sorption process is not influenced by Zn. At 26 d, a Zn shell at 3.48 Å correlates with the Cu shell at 3.40 Å in the Zn XAFS of the sample. That the Cu shell is no longer detected indicates Cu may be directly polymerizing with Zn, instead of other Cu ions, with time due to higher availability of Zn. That low Cu concentration deters Cu−Cu polymerization is evidenced by the 10:10 and 0:10 samples where Cu at higher concentration, both in the absence and presence of Zn, polymerizes with Cu ions only and forms linear bidentate complexes. At 1 d, Zn in GW solutions forms a Zn-phosphate structure that may be promoted by the presence of initial P in solution. The Zn-phosphate configuration dominates at longer reaction times because Cu sorption is delayed by complexation with organics, though there is evidence for Zn cosorption with Cu at 26 d. When struvite is precipitated from the GW solution, the O shell is more relaxed, but Zn polymers are also detected. In NN solutions, Zn forms mondentate linkages with struvite phosphate tetrahedra at 1 d, but becomes copolymerized with Cu at longer reaction times. A Zn-phosphate configuration is detectable at 7 d, but formation is limited by availability of Cu for cosorption with Zn. In turn, Cu retains direct linkages with the struvite structure and Zn polymerizes with this Cu. This mechanism may facilitate enhancement of Zn sorption by Cu as observed in batch experiments. Impact on IR-Sensitive Functional Groups. ATR-FTIR was used to determine the impact of sorbed constituents on struvite functional groups (Figure 2a). The characteristic ν3 antisymmetric stretching vibration of PO43− consists of a peaks E

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Figure 3. (a) Temperature-dependent TG, DSC, and GS curves for unreacted struvite (MAP). Data for MAP, GW, and NN solids from (b) the TG curve; (c) the DSC curve; (d) the maximum in the GS curve, and a representative 3D plot for MAP showing wavenumbers and temperatures of absorbance for evolved gases.

with growth of a broad band centered at ∼1060 cm−1. Because the fundamental difference is growth and shifting of the shoulder at ∼1060 cm−1, this is due to phenomena other than Zn or Cu interaction with struvite in GW solutions. As TOC is associated with these solids (Table 1), whether the observed differences could be attributed to organic interactions was evaluated. Experiments with model organic carbon at initial concentrations present in GW yielded lower TOC loadings, with no impact on the FTIR spectra (Table S6, Figure S4). If changes to the GW spectra are due to organics, the sorbed organics are notably different from the model organic, which would also explain differences in the extent of sorption. The SUVA value for the GW solution is 3.5 L/(mg·m); however, values for the GW solids are 11−14 L/(mg·m) (Supporting Information). Therefore, sorbed organics have higher aromaticity indicating fractionation and/or aggregation during uptake. The FTIR spectrum for the GW PPT solid is also unaffected (Figure S4), but sample preparation, and thus mechanistic differences may be a factor. Though not fully conclusive, it is feasible that changes to the ν3 PO43− band, and suppression of the ν4 NH4+ band, could indeed be attributable to organic sorption. Thermal Stability of Struvite. Thermal properties of the solids from GW and NN solutions were compared. All solids

exhibited curves characteristic of struvite (Figure 3a). In general, the dominant mass loss step in the TG, a major endothermic peak in the DSC, and maximum absorbance in the FTIR as observed in the Gram Schmidt (GS) curve, all occur within a similar temperature range for a single sample. Therefore, all solids undergo endothermic decomposition involving the loss of structural components as IR-sensitive volatiles. Though curves are similar, select thermal parameters for GW and NN solids differ from unreacted struvite (Table S7). Data selected from the TG curve shows the onset temperature at which mass loss begins, the temperature of maximum mass loss from the derivative of the TG curve (DTG), and the overall mass loss from the sample (Figure 3b). DSC results provide the temperatures of decomposition onset and the endothermic peak, with the area under the peak used to calculate the enthalpy of decomposition (Figure 3c). The temperature of these processes vary depending on solution and reaction time. Generally, temperatures shifted to lower values for GW relative to NN solids, and shifted to lower temperature with increasing reaction time and loading. Properties for the GW PPT solid were similar to those of the GW 1 d solid (Table S7). The percent mass loss from the sample in the TG and the enthalpy in the DSC also showed a similar trend, with F

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the phosphate band in the FTIR spectra even at low loadings. In the region of the phosphate band, the GW and NN 1 d samples are similar, with the shape influenced by Cu, and to a lesser extent, Zn. This suggests mechanisms of Cu sorption detected by XAFS for NN solutions are also relevant to GW solutions, for which Cu loadings are mostly below detection. Additional features in the FTIR spectra for the GW samples, affecting phosphate and ammonium groups, may be attributable to organic association with the solids. Sorption of metals and organics increase struvite susceptibility to decomposition with higher loadings enhancing the effect. In addition to the evolution of the structural constituents of struvite, NH3(g) and H2O(g), organics are also released as CO(g) and CO2(g). Implications for Sustainable Use of Struvite. Simultaneous removal of Cu, Zn, and organics with struvite recovered from GW solutions was observed. That metals cosorb and can be removed simultaneously, albeit via different mechanisms, depends on ratios, concentrations and complexing ligands. The loadings of Zn, Cu, and TOC at all reaction times and for directly precipitated struvite are below recently proposed limits for struvite fertilizer at 200 mg/kg Cu, 600 mg/kg Zn, and 2% TOC (20 000 mg/kg).44 However, higher loadings are possible if initial concentrations exceed those of the wastewater evaluated in this study. Alternatively, this material might be a good source of Cu and Zn micronutrients to agroecosystems, but consistent usage can result in metal accumulation in soils. If contaminant content is unconcerning, reduced thermal stability, a proxy for decomposition under ambient conditions, relative to contaminant-free struvite should be considered. Struvite becomes increasingly susceptible to solid-state decomposition at higher contaminant content, releasing volatiles. Evolved gases include NH3(g), a major pollutant that reacts to form aerosols and contributes to N2O greenhouse gas formation,22−24 and the greenhouse gas H2O(g).45 Organics are released as CO2(g), another greenhouse gas, and CO(g) a pollutant that also impacts climate.46 Low struvite solubility,47,48 means that release to the gas phase can be a viable pathway for structural and sorbed components that are susceptible to volatilization, albeit at slower rates under ambient conditions. This may be a primary concern for longterm storage, disposal, and reuse of the recovered mineral considering that ∼1.5 million metric tons of P is available for recovery from waste sources globally,3 and the struvite recovery process can be 70−90% efficient.1,25 Though this study was conducted for a greenhouse waste, these impacts are likely to persist in other systems where both metals and organics may interact with struvite. Results from simplified solutions are useful in the assessment of the fundamental properties of Zn and Cu sorption with struvite, and the influence of organics, as a starting point in addressing processes in complex solutions with which the mineral may interact.

thermal properties and parameters for the GW 26 d sample exhibiting the most deviation compared to both unreacted MAP and other samples. The endothermic decomposition and resultant mass loss is correlated with gas release from the sample, which results in absorbance in the FTIR. As with other parameters, the maximum absorbance in the GS curve shifts to lower temperatures from NN to GW solids, and with longer reaction times (Figure 3d). The gases released are identifiable from the FTIR spectra, and are shown as a function of temperature in the representative 3D plot for struvite (Figure 3d). The typical gases evolved are NH3(g) and H2O(g) associated with structural components of struvite. Primarily for the GW 26 d solid the temperature of maximum gas release shifts to a lower value relative to unreacted struvite from 147 to 130 °C for NH3(g) and H2O(g) (Table S8). Though the ν4 NH4+ band in the FTIR is suppressed for the GW 26 d solid, NH3(g) is evolved confirming the presence of structural ammonium. Therefore, disappearance of this band may be a surface phenomenon as a result of organic sorption as suggested above. For all GW samples, CO(g) is also evolved with maximum release shifting to lower temperature with reaction time, from 336 to 253 °C; and at 340 °C for the GW PPT solid. For the GW 26 d sample alone, CO2(g) is also released at a temperature of 405 °C. The evolution of CO(g) and CO2(g) from these samples further confirms organic association with solids from the GW solution, and susceptibility to volatilization during decomposition. Results from thermal treatment reveal the impact of sorbates on the stability of struvite. The overall trend for solids derived from both GW and NN solutions is such that with longer reaction times and higher metal concentrations the mineral becomes more susceptible to decomposition. This effect was enhanced for solids from GW solutions where organic sorption may also contribute to destabilization. Contact with aqueous solution can promote dissolution−precipitation processes that change the near-surface mineral structure, which may also impact stability. However, both NN and GW solutions reacted for similar times, especially at 26 d, show distinct effects on thermal properties, indicating that the actual constituents in solution, and the extent of sorption do have an effect. Thermal destabilization also means increased susceptibility to volatile release, which is particularly true of the GW solids for which decomposition results in evolution of CO(g) and CO2(g) along with NH3(g) and H2O(g). Uptake of Metals and Organics by Struvite in Greenhouse Wastewater. Though the GW solution is compositionally complex with several unknown constituents, simplified NN solutions were useful in elucidating processes and mechanisms. The metal loadings for GW solids were lower compared to NN solids, but could be explained by metal complexation with organics in GW solutions. The XAFS reveal some structural differences in the metal binding configuration for GW and NN solids that can be attributed to excess P, and organic complexation with Cu in the GW solution. Additional P in the GW solution may facilitate formation of a hydrated Znphosphate at short reaction times; and organic complexation delays cosorption of Cu such that this phenomenon is detectable by XAFS only at longer reaction times. In the absence of these additional constituents, Zn in NN solids forms direct linkages with struvite over short reaction times, transitioning to copolymerization with Cu. Copper forms direct linkages with struvite phosphate tetrahedra, impacting

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acssuschemeng.5b01348. Tables S1−S8, Figures S1−S4, and accompanying text: composition and speciation of, and struvite precipitation from GW; TOC, SUVA, and model DOC procedures and results; EXAFS data and extended analysis; XRD patterns; STA and EGA results (PDF). G

DOI: 10.1021/acssuschemeng.5b01348 ACS Sustainable Chem. Eng. XXXX, XXX, XXX−XXX

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AUTHOR INFORMATION

Corresponding Author

*A. A. Rouff. Email: ashaki.rouff@rutgers.edu. Phone: (973) 353-2511. Fax: (973) 353-1965. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS

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REFERENCES

Support was provided by the National Science Foundation Grant No. EAR-1506653. This research used the Advanced Photon Source, Argonne National Laboratory operated under Contract No. DE-AC02-06CH11357. Thanks to S. Lee and B. Reinhart of Beamline 12BM-B for technical support, and to N. Ma, PhD for assistance with preparation of GW PPT.

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DOI: 10.1021/acssuschemeng.5b01348 ACS Sustainable Chem. Eng. XXXX, XXX, XXX−XXX