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Comparison of particulate pyrite autotrophic denitrification (PPAD) and sulfur oxidizing denitrification (SOD) for treatment of nitrified wastewater Shuang Tong, Laura C. Rodriguez-Gonzalez, Chuanping Feng and Sarina J. Ergas
ABSTRACT The use of reduced sulfur compounds as electron donors for biological denitrification has the potential to reduce chemical and sludge disposal costs as well as carry-over of organic carbon to the effluent that often occurs with heterotrophic denitrification. Although a number of prior studies have evaluated sulfur oxidizing denitrification (SOD), no prior studies have evaluated particulate pyrite autotrophic denitrification (PPAD) in continuous flow systems. Bench-scale upflow packed bed reactors (PBRs) were set up to compare denitrification rates, by-product production and alkalinity consumption of PPAD and SOD. At an empty bed contact time of 2.9 h, average NO 3 -N removal efficiencies were 39.7% and 99.9% for PPAD and SOD, respectively. Although lower denitrification rates were observed with PPAD than SOD, lower alkalinity consumption and reduced sulfur 2 2 by-product formation (SO2 and SO2 4 , S 3 plus S2O3 ) were observed with PPAD. Furthermore,
Shuang Tong Chuanping Feng School of Water Resources and Environment, China University of Geosciences Beijing, Beijing 100083, China Shuang Tong Laura C. Rodriguez-Gonzalez Sarina J. Ergas (corresponding author) Department of Civil and Environmental Engineering, University of South Florida, 4202 E. Fowler Ave, ENB 118, Tampa, FL 33620, USA E-mail:
[email protected]
higher denitrification rates and lower by-product production was observed for SOD than in prior studies, possibly due to the media composition, which included sand and oyster shells. The results show that both pyrite and elemental sulfur can be used as electron donors for wastewater denitrification in PBRs. Key words
| biological nitrogen removal (BNR), NO 3 removal, particulate pyrite autotrophic denitrification (PPAD), sulfur oxidizing denitrification (SOD), wastewater treatment
INTRODUCTION Biological denitrification can be used to remove nitrate (NO 3 ) from nitrified domestic, industrial and agricultural wastewaters (Li et al. ), contaminated drinking water (Moon et al. ) and storm water runoff (Borne et al. ). Autotrophic denitrification is carried out by chemolithotrophic bacteria that couple oxidation of either hydrogen (H2), reduced sulfur or reduced iron compounds with NO 3 reduction under anoxic conditions. Advantages of autotrophic denitrification over heterotrophic denitrification are that no external organic carbon source is required and there is low excess biomass (e.g. sludge) production or carry-over of organic carbon to the effluent (Sengupta et al. ). Conventional sulfur oxidizing denitrification (SOD), which utilizes elemental sulfur (S0) as an electron donor, is the most common sulfur-related autotrophic denitrification technology. The following stoichiometric doi: 10.2166/wst.2016.502
equation describes the SOD process (Batchelor & Lawrence ): þ 1:10S0 þ 0:40CO2 þ NO 3 þ 0:76H2 O þ 0:08NH4 þ ! 0:08C5 H7 O2 N þ 0:50N2 þ 1:10SO2 4 þ 1:28H
(1)
Although SOD has several advantages over heterotrophic denitrification, its main disadvantages are sulfate (SO2 4 ) production and alkalinity consumption (Lee et al. ). Based on Equation (1), for every 1.0 g of NO 3 -N removed by SOD, 2.51 g of SO2 -S are produced and 4.57 g 4 of alkalinity (as CaCO3) are consumed. The US Environmental Protection Agency (USEPA) guideline for SO2 4 in drinking water is 83 mg S/L (250 mg/L as SO2 ). In addition, 4 the optimum pH for sulfur oxidizing bacteria, such as Thiobacillus denitrificans, is 6.0–8.0 (Holt et al. ), which can be difficult to obtain in SOD of low alkalinity waters. In light of
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the alkalinity consumption, an alkalinity source is required for SOD. Sengupta et al. () investigated solid-phase buffers, marble chips, limestone, and crushed oyster shells for SOD. They reported that crushed oyster shell was the most suitable buffer for SOD based on the ability of the buffer surface to act as host for microbial attachment and the slow dissolution rate of the buffering agent (CaCO3) released. Although most prior studies of SOD do not include biofilm carriers, Krayzelova et al. () included scrap tire chips in a SOD column and reported high denitrification rates and low SO2 production even under highly transient loading 4 conditions. Iron-sulfur minerals are abundant in the earth’s crust and have attracted attention as potential electron donors for autotrophic denitrification. Li et al. () found that chemical grade FeS was an effective substrate for denitrification in batch experiments carried out under varying pH, temperature and NO 3 concentrations. In further studies, pyrrhotite, an abundant mineral (Fe1-xS, where x varies from 0 to 0.125), was used as an electron donor for denitrification in packed bed reactors (PBRs) and achieved 77% removal of NO 3 at a 12 h empty bed contact time (EBCT) (Li et al. ). Pyrite (FeS2), has also been studied as a potential denitrification substrate. Otero et al. () coupled NO 3 and sulfate isotopic data and showed that denitrification occurred with pyrite oxidation in samples collected from groundwater flowing through an aquifer with a pyrite mine. Using the method of McCarty () and thermodynamic data from David (), a stoichiometric equation for pyrite oxidizing denitrification was calculated as: þ 0:364FeS2 þ 0:116CO2 þ NO 3 þ 0:821H2 O þ 0:023NH4 ! 0:364 FeðOHÞ3 þ0:023C5 H7 O2 N þ 0:50N2 þ þ 0:729SO2 4 þ 0:480H
Table 1
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(2)
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The low biomass yield predicted (0.085 g VSS/g NO 3 -N; VSS: volatile suspended solids) compared with SOD (0.13 g VSS/g NO 3 -N) is due to the low Gibbs free energy change for the reaction of pyrite and nitrate (34.13 kJ/eeq transferred) compared with SOD (53.05 kJ/eeq transferred). Based on Equation (2), for every 1.0 g of NO 3 -N removed, 1.67 g of SO2 4 -S are produced and 1.71 g of alkalinity (as CaCO3) are consumed. Pu et al. () reported that pyrite was a good electron donor for denitrification and observed that the pH was stable in pyrite-based autotrophic denitrification microcosms. However, no prior studies have investigated particulate pyrite autotrophic denitrification (PPAD) in a continuous flow bioreactor system for treatment of nitrified wastewater. The overall goal of this research was to compare PPAD and SOD performance for treatment of nitrified domestic wastewater in side-by-side column studies. The medium used in these columns included crushed oyster shells as a solid phase buffer and biofilm carrier and quartz sand as an additional biofilm carrier. The specific objectives were to: (1) compare the denitrification performance of PPAD and SOD columns, (2) investigate by-product formation in both PPAD and SOD columns and (3) evaluate SOD in PBRs with a medium that includes oyster shells and sand.
MATERIALS AND METHODS The column study was divided into two phases, as shown in Table 1. Phase I was set up to compare PPAD and SOD performance. In Phase II, the effects of increased pyrite dose and increased flow rate (decreased EBCT) were investigated. PPAD and SOD column study Two 750 mL acrylic columns (height ¼ 35 cm; inside diameter ¼ 5.2 cm) were set up as upflow PBRs at room
Description of experimental phases
Media composition (g) Phase
Study length (d)
Flow rate (L/d)
EBCT (h)
Column
P
S0
OS
Sa
I
38
1.0
3.5
PPAD SOD
400 –
– 350
400 400
500 280
II
30
1.2
2.9
PPAD SOD
800 –
– 350
400 400
390 280
P, particulate pyrite; S0, sulfur pellets; OS, oyster shells, S, sand. a
Sand was used to fill the remainder of the column.
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temperature (20 ± 2 C). Each column was equipped with five sample ports along the length, in addition to inlet and outlet ports. The columns were filled with different amounts of particulate pyrite (P; 1–2 mm), elemental sulfur pellets (S0; 2.5–3.5 mm), crushed oyster shell (OS; 50% of 1–2 mm; 50% of 2–4 mm) and quartz sand (S; 1–2 mm) during different phases, as shown in Table 1. The initial amount of FeS2 and S0 (Phase I) was based on the stoichiometry of PPAD (Equation (2)) and SOD (Equation (1)), assuming an influent NO 3 -N concentration of 100 mg/L and the substrate requirements for an operating period of 3 years. The values obtained were then multiplied by a safety factor of 1.3. The amount of oyster shell added to both columns was calculated based on the alkalinity required for SOD (Equation (1)) multiplied by a safety factor of 1.3. Note that a 3 year operational period could only be achieved if periodic maintenance were performed (e.g. backwashing) to remove excess biomass and mineral precipitates. The columns were seeded with mixed liquor suspended solids from the Falkenburg Advanced Water Treatment facility in Hillsborough County, FL, USA, to enrich PPAD and SOD microbial communities. Both columns were operated with a closed recirculation system for the first 8 days to enhance attachment and acclimation of the biofilm. After the 8 d start-up period, a 30 d steady flow period was run as Phase I. Based on the Phase I results, the wastewater flow rate was increased for both columns and additional particulate pyrite was added to the PPAD column in Phase II (Table 1). W
Wastewater composition Experiments were carried out using synthetic nitrified wastewater with the following composition (per litre of deionized water): NaHCO3 (900 mg), KNO3 (721 mg), KH2PO4 (44 mg), NH4Cl (38 mg). This composition resulted in an influent NO 3 -N concentration of 100 mg/L and an influent alkalinity of 290 ± 5 mg/L as CaCO3. Note that no SO2 4 was included in the synthetic wastewater so that all effluent 2 SO2 4 would be due to SO4 formation in the process. Analytical methods Sample collection was performed daily for the effluent from both PPAD and SOD columns. Influent samples were collected approximately three times a week, whenever a fresh batch of influent wastewater was prepared. Samples were filtered through 0.45 μm membrane filters.
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2 NO and NHþ 3 , NO2 , SO4 4 concentrations were measured in the filtrate by ion chromatography (USEPA ) using an 881 Compact IC pro system (Metrohm AG, Switzerland). Method detection limits (MDLs) for NO 3 -N, 2 þ NO2 -N, SO4 -S and NH4 -N were 0.01, 0.04, 0.01 and 0.20 mg/L, respectively. The H2O2 oxidation method was used to detect SO2 and S2O2 (Pu et al. ). In this 3 3 2 method, SO3 and S2O2 were oxidized to SO2 3 4 and then 2 the SO4 was measured by ion chromatography (Pu et al. ). Standard methods (APHA ) were used to measure concentrations of S2, FeTotal, Fe2þ and alkalinity (methods 4500 S2 D; 3500-Fe B; 2320 B). The MDLs were 0.005, 0.004 and 0.004, 20 mg/L, respectively. The pH was measured using a pH meter and calibrated electrode (Orion 5 Star, Thermo Scientific Inc., USA; MDL: 0–14).
RESULTS AND DISCUSSION Phase I PPAD and SOD performance Phase I of the column study was carried out to compare the performance of PPAD and SOD when the columns were packed with equivalent amounts of electron donor based on the stoichiometry shown in Equations (1) and (2). After an initial start-up period of 8 d with continuous recirculation, biofilm formation and significant NO 3 removal were observed in both columns (Figure 1). After the start-up period, the columns were switched to continuous flow mode (Q ¼ 1.0 L/d; EBCT ¼ 3.5 h). Average effluent NO 3 -N concentrations were 43.29 ± 7.34 and 0.31 ± 0.49 mg/L for PPAD and SOD columns, respectively (Table 2). The average NO 3 removal efficiencies were approximately 56.4% and 99.7% for the PPAD and SOD columns, respectively. þ Effluent NO 2 -N and NH4 -N concentrations are also þ shown in Figure 1. Both NO 2 -N and NH4 -N concentrations in the SOD column were lower than in the PPAD column. NHþ 4 removal was likely due to biosynthesis and/or nitrification in aerobic zones near the inlet. Li et al. () found that NHþ 4 -N concentrations up to 7.8 mg/L positively influenced denitrification rates when FeS was utilized as a denitrification substrate, most likely due to the preference for NHþ 4 as a nitrogen source for biosynthesis. NO production was 2 likely due to partial nitrification and/or denitrification. Average effluent secondary S by-product concentrations 2 2 2 (SO2 4 , S , SO3 plus S2O3 ) are shown in Table 3. Aver2 age effluent SO4 -S concentrations for the PPAD column were lower than the USEPA drinking water guideline
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Figure 1
Table 2
Comparison of pyrite and sulfur oxidizing denitrification
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2 Concentrations of NO3 -N, NO2 -N, NHþ 4 -N and SO4 -S in the influent and effluent in the PPAD and SOD columns.
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Summary of average N species concentrations in the influent and effluent of PPAD and SOD during Phases I and II
N species concentrations Experimental phases
NO3 -N mg/L
NO2 -N mg/L
NHþ 4 -N mg/L
TIN %
NO3 -N RE%
TIN RE%
Phase I Influent
99.28 ± 1.70
–
10.30 ± 0.40
109.58
–
–
PPAD
43.29 ± 7.34
8.09 ± 4.91
5.21 ± 0.55
56.59
56.4
48.4
SOD
0.31 ± 0.49
3.15 ± 1.97
2.81 ± 0.67
6.27
99.7
94.3
Phase II Influent
99.67 ± 0.73
–
10.40 ± 0.41
110.07
–
–
PPAD
60.08 ± 4.80
1.87 ± 1.19
7.93 ± 0.82
69.88
39.7
36.5
SOD
0.09 ± 0.07
2.56 ± 0.66
3.83 ± 1.60
6.48
99.9
94.1
TIN , total inorganic nitrogen; RE, removal efficiency.
(83 mg/L), and SO2 4 -S productivity was only 0.12 ± 0.06 mg 2 SO4 -S/mg NO3 -N, which was lower than the value pre dicted by Equation (2) of 0.73 mg SO2 4 -S/mg NO3 -N. In 2 2 2 addition, concentrations of S and SO3 plus S2O3 were below the MDL in the PPAD column (Table 3). FeTotal and Fe2þ were also measured in the effluent, as Fe is one
of the constituent elements of pyrite. However, FeTotal and Fe2þ concentrations were below the MDL in both columns. The low production of pyrite oxidation by-products in the PPAD column was likely due to accumulation of intracellular S0 storage products and/or precipitation reactions (Miotlin´ski ; Zhang et al. ; Li et al. ). These
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Table 3
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Summary of average concentrations of S species, alkalinity and pH in the influent and effluent of PPAD and SOD during Phases I and II
S species concentrations mg/L
Experimental phases
SO24 -S
Others
SO23 -S plus S2O23 -S
Alkalinity
pH
S2
mg CaCO3/L
–
Phase I Influent
–
–
–
290 ± 5
7.75 ± 0.08
PPAD
6.81 ± 3.45
BDL
BDL
360 ± 20
7.72 ± 0.13
SOD
166.13 ± 18.10
10.60 ± 1.25
14.15 ± 2.83
290 ± 20
6.94 ± 0.16
Phase II Influent
–
–
–
290 ± 5
7.70 ± 0.02
PPAD
3.94 ± 1.15
BDL
BDL
320 ± 40
7.80 ± 0.15
SOD
200.37 ± 33.51
9.75 ± 0.91
30.99 ± 2.60
265 ± 20
6.75 ± 0.26
BDL, below MDL.
mechanisms are currently being investigated further in our laboratory. The average effluent SO2 4 -S concentration in the SOD column was higher than the EPA drinking water guideline (Table 3). However, the average SO2 4 -S productivity of 1.66 ± 1.81 mg SO2 4 -S/mg NO3 -N in the SOD column was lower than the stoichiometric value (Equation (1)) of 2.51 mg SO2 4 -S/mg NO3 -N and was also lower than observed prior studies (Zhang et al. ). The presence of S2 in the effluent (Table 3) was likely due to sulfur dispro2 portionation (S0 ! SO2 reduction in the 4 þ H2S) or SO4 upper sections of the column where NO 3 is depleted (Boles et al. ). Although the synthetic wastewater did not contain any added organic carbon source, the carbon source for SO2 reduction could come from the oyster 4 shells or sloughed biofilm. Both alkalinity concentration and pH in the PPAD column effluent (360 ± 20 mg CaCO3/L; 7.72 ± 0.13) were higher than in the SOD column (290 ± 20 mg CaCO3/L; 6.94 ± 0.16) (Table 3), showing that the PPAD process consumed less alkalinity. However, both columns maintained a near-neutral pH in the range that has been shown to promote SOD (Holt et al. ). 2 Profiles of NO 3 -N and SO4 -S concentrations vs. depth from inlet for both columns during Phase I are shown in Figure 2. Overall NO 3 -N removal rates were calculated by subtracting the effluent concentration from the influent and dividing by the EBCT. Average overall NO 3 -N removal rates during Phase I were 61.9 and 120.0 mg/(L·d) for PPAD and SOD, respectively. Lower removal rates in the PPAD columns may have been due to biokinetic factors or a slower dissolution rate of the electron donor from pyrite than from the S0 pellets. NO 3 concentration profiles for
the SOD column were fit to a first order equation (k ¼ 1.75 h1; R² ¼ 0.97): C ¼ Co expðkτ Þ
(3)
where Co is the influent NO 3 -N concentration (mg/L), k is the pseudo-first order rate coefficient (h1) and τ is EBCT (h). Phase II PPAD and SOD system performance During Phase II the EBCT was decreased to 2.9 h in both columns by increasing the wastewater flow rate to 1.2 L/d. This EBCT was set to achieve a predicted effluent NO 3 -N concentration of ∼1 mg/L for the SOD column based on Equation (3). The mass of particulate pyrite was increased to 800 g in the PPAD column to increase the surface area available for biofilm formation based on preliminary studies showing increased denitrification with increased pyrite dose (data not shown). The average overall NO 3 -N removal rate in the PPAD column (56.2 mg/(L·d)) was approximately the same as that of Phase I (Figures 2(a) and 3(a)). Although this rate was much lower than for SOD in this study, the rate was similar to prior published studies of SOD (Sahinkaya & Dursun ). Effluent NO 2 -N concentrations decreased substantially in Phase II, from 8.09 ± 4.91 mg/L to 1.87 ± 1.19 mg/L (Figure 1; Table 2), indicating that more complete denitrification occurred, due either to the higher pyrite dosage or acclimation of the biofilm. NO 2 is recognized for its toxic effects on both human and animal health and results in methemoglobinemia (Chen et al. ). Hence, increasing the amount of particulate pyrite can
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Figure 2
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Profiles of NO3 -N and SO24 -S concentrations vs. depth from inlet in Phase I. Due to the low SO24 productivity in the PPAD column the vertical axis is expanded in the inset of (b).
decrease the accumulation of by-products of incomplete denitrification. NO 3 was effectively removed in the SOD column at the higher flow rate in Phase II (Figure 1 and Table 2). The results fit a first order equation (Figure 3), with a pseudo-
Figure 3
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Profiles of NO3 -N and SO24 -S concentrations vs. depth from inlet in Phase II.
first order rate coefficient (k ¼ 1.48 h1; R2 ¼ 0.98) that was lower than in Phase I. A number of factors affect mass transfer and biodegradation in PBRs, including biofilm thickness and fluid velocity, which could have influenced the rate coefficient.
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The overall NO 3 -N removal rate (820 mg/(L·d)) was higher and the SO2 4 -S productivity was lower (2.01 mg SO2 4 -S/mg NO3 -N) in this study than has been observed in prior studies of SOD (Sahinkaya & Dursun ; Smith ). This improved performance was most likely due to the packing medium used, which included elemental sulfur pellets, quartz sand and crushed oyster shells. Quartz sand and oyster shells have a higher specific surface area than S0 pellets, resulting in increased available surface area for biofilm attachment (Sengupta et al. ; Conneely ). In addition, oyster shells consist of a hard tissue of calcium carbonate in an organic matrix, which has been identified as protein and carbohydrate (Simkiss ). Several previous studies have shown that the addition of an organic substrate enhances denitrification performance of SOD and reduces sulfur by-product production (Sahinkaya & Dursun ; Smith ; Krayzelova et al. ). In addition, Li et al. () observed lower SO2 4 -S productivity in PBRs packed with a mixture of limestone and pyrrhotite compared with PBRs packed with pyrrhotite alone.
CONCLUSIONS Column studies were carried out to compare and optimize the performance of PPAD and SOD for treatment of nitrified wastewater in packed bed bioreactors. At an EBCT of 2.9 h, average NO 3 -N removal efficiencies were 39.7% and 99.9% for PPAD and SOD, respectively. The results of this study show that a PBR medium that included S0 pellets, crushed oyster shells and quartz sand resulted in higher denitrification rates and lower by-product production than found by prior studies. Although denitrification rates were lower in the PPAD column, lower alkalinity consumption and reduced by-product production were observed compared with SOD. This study is the first to utilize particulate pyrite and an electron donor for denitrification in continuous flow packed bed bioreactors. Additional research is needed to understand the mechanisms of denitrification and to enhance the rate of denitrification in PPAD.
ACKNOWLEDGEMENTS The authors acknowledge Justine Stocks, a graduate student at the University of South Florida for her assistance in sample collection. This research work was supported by China Scholarship Council (CSC, No. 201406400007) during a visit of Shuang Tong to University of South Florida.
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This publication was also made possible by USEPA grant 83556901. Its contents are solely the responsibility of the grantee and do not necessarily represent the official views of the USEPA. Further, USEPA does not endorse the purchase of any commercial products or services mentioned in the publication.
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treatment: elimination of excess sulfate production and alkalinity requirement. Chemosphere 89, 144–149. Sengupta, S., Ergas, S. J. & Lopez-Luna, E. Investigation of solidphase buffers for sulfur-oxidizing autotrophic denitrification. Water Environment Research 79 (13), 2519–2526. Simkiss, K. The organic matrix of the oyster shell. Comparative Biochemistry and Physiology 16 (4), 427–435. Smith, D. P. Onsite wastewater nitrogen reduction with expanded media and elemental sulfur biofiltration. Water Science and Technology 65 (4), 750–756. USEPA Methods for the Determination of Organic and Inorganic Compounds in Drinking Water. United States Environmental Protection Agency. EPA/815-R-00-014, Washington, DC. Zhang, Y.-C., Slomp, C. P., Broers, H. P., Passier, H. F. & Van Cappellen, P. Denitrification coupled to pyrite oxidation and changes in groundwater quality in a shallow sandy aquifer. Geochimica et Cosmochimica Acta 73 (22), 6716–6726. Zhang, L., Zhang, C., Hu, C., Liu, H. & Qu, J. Denitrification of groundwater using a sulfur-oxidizing autotrophic denitrifying anaerobic fluidized-bed MBR: performance and bacterial community structure. Applied Microbiology and Biotechnology 99 (6), 2815–2827.
First received 16 August 2016; accepted in revised form 8 October 2016. Available online 8 November 2016
