Scholarly article on topic 'Membrane biological reactor treatment of a saline backwash flow from a recirculating aquaculture system'

Membrane biological reactor treatment of a saline backwash flow from a recirculating aquaculture system Academic research paper on "Environmental engineering"

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Aquacultural Engineering
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{"Recirculating system" / "Effluent treatment" / "Waste capture" / "Membrane biological reactor" / Salinity / "Water reclamation"}

Abstract of research paper on Environmental engineering, author of scientific article — Mark J. Sharrer, Yossi Tal, Drew Ferrier, Joseph A. Hankins, Steven T. Summerfelt

Abstract A recirculating aquaculture system (RAS) can minimize water use, allowing fish production in regions where water is scarce and also placing the waterborne wastes into a concentrated and relatively small volume of effluent. The RAS effluent generated during clarifier backwash is usually small in volume (possibly 0.2–0.5% of the total recirculating flow when microscreen filters are used) but contains high levels of concentrated organic solids and nutrients. When a RAS is operated at high salinities for culture of marine species, recovering the saltwater contained in the backwash effluent could allow for its reuse within the RAS and also reduce salt discharge to the environment. Membrane biological reactors (MBRs) combine activated sludge type treatment with membrane filtration. Therefore, in addition to removing biodegradable organics, suspended solids, and nutrients such as nitrogen and phosphorus, MBRs retain high concentrations of microorganisms and, when operated with membrane pore sizes <1μm, exclude microorganisms from their discharge. In this research, an Enviroquip (Austin, TX) MBR pilot-plant was installed and evaluated over a range of salinities to determine its effectiveness at removing bacteria, turbidity, suspended solids, nitrogen, phosphorus and cBOD5 content from the approximately 22m3/day concentrated biosolids backwash flow discharged from the RASs at The Conservation Fund Freshwater Institute. The MBR system was managed at a hydraulic retention time of 40.8h, a solids retention time of 64±8 days, resulting in a Food: Microorganism ratio of 0.029day−1. Results indicated excellent removal efficiency (%) of TSS (99.65±0.1 to 99.98±0.01) and TVS (99.96±0.01 to 99.99±0.0) at all salinity levels. Similarly, a 3–4log10 removal of total heterotrophic microbes and total coliform was seen at all treatment conditions. Total nitrogen removal efficiency (%) ranged from 91.8±2.9 to 95.5±0.6 at the treatment levels and was consistent, provided a sufficient acclimation period to each new condition was given. Conversely, total phosphorus removal efficiencies (%) at 0ppt, 8ppt, 16ppt and 32ppt salinity were 96.1±1.0, 72.7±3.5, 70.4±2.3, and 65.2±5.4, respectively, indicating reduced phosphorus removal at higher salinities.

Academic research paper on topic "Membrane biological reactor treatment of a saline backwash flow from a recirculating aquaculture system"

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Aquacultural Engineering 36 (2007) 159-176

Membrane biological reactor treatment of a saline backwash flow from a recirculating aquaculture system

Mark J. Sharrera, Yossi Talb, Drew Ferrierc, Joseph A. Hankinsa,

Steven T. Summerfelta'*

a The Conservation Fund's Freshwater Institute, 1098 Turner Road, Shepherdstown, WV 25443, United States b Center of Marine Biotechnology, University of Maryland Biotechnology Institute, 701 E. Pratt Street, Baltimore, MD 21202, United States c Hood College, Department of Environmental Biology, 401 Rosemont Avenue, Frederick, MD 21701-8575, United States

Received 30 May 2006; accepted 16 October 2006


A recirculating aquaculture system (RAS) can minimize water use, allowing fish production in regions where water is scarce and also placing the waterborne wastes into a concentrated and relatively small volume of effluent. The RAS effluent generated during clarifier backwash is usually small in volume (possibly 0.2-0.5% of the total recirculating flow when microscreen filters are used) but contains high levels of concentrated organic solids and nutrients. When a RAS is operated at high salinities for culture of marine species, recovering the saltwater contained in the backwash effluent could allow for its reuse within the RAS and also reduce salt discharge to the environment. Membrane biological reactors (MBRs) combine activated sludge type treatment with membrane filtration. Therefore, in addition to removing biodegradable organics, suspended solids, and nutrients such as nitrogen and phosphorus, MBRs retain high concentrations of microorganisms and, when operated with membrane pore sizes <1 mm, exclude microorganisms from their discharge. In this research, an Enviroquip (Austin, TX) MBR pilot-plant was installed and evaluated over a range of salinities to determine its effectiveness at removing bacteria, turbidity, suspended solids, nitrogen, phosphorus and cBOD5 content from the approximately 22 m3/day concentrated biosolids backwash flow discharged from the RASs at The Conservation Fund Freshwater Institute. The MBR system was managed at a hydraulic retention time of 40.8 h, a solids retention time of 64 ± 8 days, resulting in a Food: Microorganism ratio of 0.029 day-1. Results indicated excellent removal efficiency (%) of TSS (99.65 ± 0.1 to 99.98 ± 0.01) and TVS (99.96 ± 0.01 to 99.99 ± 0.0) at all salinity levels. Similarly, a 3-4 logi0 removal of total heterotrophic microbes and total coliform was seen at all treatment conditions. Total nitrogen removal efficiency (%) ranged from 91.8 ± 2.9 to 95.5 ± 0.6 at the treatment levels and was consistent, provided a sufficient acclimation period to each new condition was given. Conversely, total phosphorus removal efficiencies (%) at 0ppt, 8 ppt, 16ppt and 32ppt salinity were 96.1 ± 1.0, 72.7 ± 3.5, 70.4 ± 2.3, and 65.2 ± 5.4, respectively, indicating reduced phosphorus removal at higher salinities. # 2006 Elsevier B.V. All rights reserved.

Keywords: Recirculating system; Effluent treatment; Waste capture; Membrane biological reactor; Salinity; Water reclamation

* Corresponding author. Tel.: +1 304 876 2815; fax: +1 304 870 2208.

E-mail addresses: (M.J. Sharrer), (S.T. Summerfelt).

1. Introduction

1.1. Background

As the global population continues its exponential rise, the demands placed on natural resources are increasing. Technologies aimed at maximizing food

0144-8609/$ - see front matter # 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.aquaeng.2006.10.003

production capabilities, environmental compatibility, and profitability are continually being developed. Agricultural practices and expertise have been expanded to allow for higher yields and lower incidence of disease. Similarly, the field of aquaculture has aspired to develop progressively more sustainable, efficient, and economical production capabilities. And, as yields of marine fish continue to decline, fish production at aquaculture facilities is becoming progressively more important. Although production in these facilities is rising, challenges associated with the intensification of this production method are ubiquitous. These issues can range from maintaining proper water quality, mechanical maintenance of production equipment, and controlling outbreak of disease. Another key issue that is encountered with the intensification of fish culture systems is effective waste management and disposal.

Water usage in fish culture facilities ranges from low exchange ponds, to complete flow-through systems, to tank-based systems using water recirculating technologies. Daily flows emanating from fish farms coupled with cleaning events that are performed to reduce suspended solids and improve water quality within an aquaculture system can result in significant discharge of waste material (Summerfelt, 1999). Components of waste resulting from fish production include nitrogen and phosphorus compounds, suspended solids, biochemical oxygen demand, and bacteria. One of the benefits of recirculating aquaculture systems is their capacity to concentrate the particulate waste materials into a relatively small waste stream. Wastewater reclamation is especially significant when marine species are being raised within systems that treat and recirculate brackish or full-strength seawater at inland locations because discharge of the salts to a freshwater watershed could be regulated and can also increase the fish farm's variable costs.

Semi-closed recirculating systems must flush the concentrated biosolids contained in filter backwash flows. The biosolids in the backwash flows are then thickened (Chen et al., 1997; Ebeling et al., 2003, 2006; Brazil and Summerfelt, 2006; Summerfelt et al., 1999) and the resulting supernatant or filter permeate often requires further treatment (Brazil and Summerfelt, 2006; Ebeling et al., 2003) and could potentially be reclaimed in order to reuse its water, salts, or heat. Further treatment of the thickened sludge involves long term storage, composting, and land application (Chen et al., 1997; Summerfelt, 1999; Summerfelt et al., 1999). The objective of this paper is to evaluate a membrane biological filtration system for reclaiming water, salts, and heat found

within the backwash flow discharged from semi-closed recirculating aquaculture systems.

1.2. Membrane filtration

A recent advancement in waste treatment technology involves the filtration of wastewater through porous membranes. Specifically, membrane biological reactors (MBRs) combine the activated sludge process of a conventional activated sludge (CAS) system with a membrane submerged in the process water capable of filtering particulate waste constituents from the mixed liquor solution. This semi-permeable membrane can retain particles greater than 0.01-10 mm, depending upon pore size, while allowing dissolved components and water to pass through the membrane (Viadero and Noblet, 2002). The liquid that passes through the membrane is referred to as permeate while the liquid excluded by the membrane is known as retentate (Crites and Tchobanoglous, 1998). As a result, components of wastewater such as suspended solids, microorganisms, and bacteria, along with the associated particulate nitrogenous components, biological oxygen demand (cBOD5), and chemical oxygen demand (COD) can be selectively excluded from the effluent of MBRs (Gunder, 2001). Membrane filtration that falls within the category of micro-filtration (pore size 0.1-10 mm) has shown the potential for pre-treatment of drinking water by removing colloidal particles, microorganisms, and other particulate material (Van der Bruggen et al., 2003). Similarly, membrane filtration has been used for surface water treatment in the Los Angeles area resulting in permeate turbidity of <0.1 ntu (Karimi et al., 2002). Membrane biological reactors have been shown to take municipal wastewater flows and after treatment provide high quality, reusable, particle free effluent (DiGiano et al., 2004; Fleischer et al., 2005; Marrot et al., 2004; Churchouse, 2001; Churchouse and Wildgoose, 1999). Consequently, treatment of the backwash flows produced in marine recirculating aquaculture systems with MBRs can potentially reclaim the water and its salt and heat for reuse in the fish production systems, while simultaneously reducing salt discharge to the environment.

Through the activated sludge process, using a recirculating loop that includes anoxic and aerobic treatment basins coupled with membrane filtration, an environment is created that is suitable for the removal of nitrogen from the wastewater through the mechanisms of nitrification and denitrification. Nitrification, which is a two-stage process and takes place in an aerobic environment, occurs when un-ionized ammonia (NH3)

is oxidized to nitrite (NO2 ) (Eq. (1)), which is further oxidized to nitrate (NO3~) (Eq. (2)):

NH3 + 1.5O2 $ NO2~ + H2O + H+ + 84 kcal moP1

NO2~ + 0.5O2 $ NO3" + 17.8 kcal moP1 (2)

A community of autotrophic bacteria that utilize free NH3 molecules and NO2~ ions as energy sources facilitates this microbiological process. Nitrosomonas spp. and Nitrobacter spp., respectively, perform this sequential action, and are cultivated within the mixed liquor suspended solids contained in the membrane filtration system (Hagopian and Riley, 1998).

Biological nitrate removal can be accomplished through either dissimilatory or assimilatory pathways (EPA, 1993a; van Rijn et al., 2006). Denitrification occurs in one of two possible dissimilatory pathways in which nitrate ions resulting from nitrification are then available for reduction to nitrogen gas by facultative anaerobes under anoxic conditions (Stephenson et al., 2000; van Rijn et al., 1995). In the second dissimilatory pathway nitrate is reduced to ammonia by obligate and facultative anaerobes under anoxic conditions; thus, both processes result in concomitant release of energy used by the bacteria. Denitrifying bacteria utilize nitrate, in the same way as oxygen, as electron acceptors and organic carbon usually serves as an electron source (EPA, 1993a; Brazil, 2004; van Rijn et al., 2006). The stoichoimetric relationship of the denitrification process is described in the following unbalanced equation (Eq. (3)) (EPA, 1993a):

NO3~ + CH3OH + H2CO3 ! N2 + H2O + HCO3"

Denitrification can also occur where facultative anaerobes reduce NO2~ to elemental nitrogen (N2) (e.g., (4)), which produces the intermediate compounds nitric oxide (NO) and nitrous oxide (N2O) under certain conditions (EPA, 1993a; van Rijn et al., 2006):

NO3~ ! NO2~ ! NO ! N2O ! N2 (4)

Finally, the assimilatory pathway occurs when microorganisms utilize nitrate to produce ammonia, which is then utilized as a nitrogen source to generate biomass (Eq. (5)) (EPA, 1993a; van Rijn et al., 2006; Brazil, 2004):

NO3~ ! NO2~ ! NH4+ (5)

Another ammonia oxidation mechanism found in urban estuarine sediments and known to be coupled with wastewater treatment technology, is anaerobic ammonia oxidation or anammox (Tal et al., 2005). These autotrophic bacteria, which use nitrite as the preferred electron acceptor and CO2 as a carbon source, catalyze this reaction according to the following equation (Tal et al., 2004):

NH3 + HNO2 ! N2 + 2H2O (6)

Conditions maintained within the membrane biological reactor likely cultivate the organisms capable of performing this microbiological mechanism as well.

Denitrification can occur in a traditional activated sludge process using an aerobic bioreactor combined with a digestion basin kept under anoxic conditions (Aboutboul et al., 1995). A wastewater treatment plant utilizing an anoxic/oxic concept showed a 99.9% reduction in NO3-N (Beeman and Reitberger, 2003). In a study by Sadick et al. (1996) that analyzed the performance of an anaerobic fluidized bed bioreactor, microorganisms attached to the suspended sand particles reduced the nitrate (NO3) concentration from 7.2 mg/L at the inlet to 0.3 mg/L in the effluent. In typical membrane bioreactor systems, the aerated and anoxic components of the coupled nitrification and de-nitrification processes are connected with a pump that recycles water from the anoxic to aerobic tank. The membrane component is located in the aerobic tank to take advantage of aeration used to scour solids from the membrane. An overflow drain from the aerobic tank to the anoxic tank maintains a constant wastewater level in the aerobic tank.

Phosphorus removal can also be accomplished within the MBR process simultaneously with nitrifica-tion/denitrification. The mechanism of phosphorus removal is both biological and physical. Phosphorus is an essential nutrient utilized by microorganisms for cell synthesis, maintenance, and energy transport (EPA, 1993b). The phosphorus accumulated by heterotrophic bacteria within the activated sludge is subsequently retained by the MBR when bacteria is excluded from the permeate flow. Enhanced biological phosphorus removal (EBPR) by de-nitrifying bacteria in the activated sludge process is realized by subjecting the mixed liquor suspended solids to alternating aerobic and anaerobic conditions (EPA, 1993b). In the anaerobic stage, phosphorus is released from the bacterial biomass. Subsequently, luxury uptake of phosphorus by microorganisms occurs in a vigorously aerated and mixed aerobic zone of this sequential process (Crites and Tchobanoglous, 1998; Barak and

van Rijn, 2000; EPA, 1993b). An alternate mechanism shows that, in an anaerobic environment, polyphosphate accumulating organisms (PAOs) convert acetate to polyhydroxyalkanoates (PHA), with simultaneous degradation of polyphosphate and release of phosphate (H3PO4) (Barak et al., 2003). Then, under anoxic conditions, phosphate is incorporated into cellular mass and polyphosphate is produced intracellularly (Barak et al., 2003).

1.3. General experiences MBR systems

Although MBRs in wastewater treatment are a relatively new tool, their application is rapidly increasing. In year 2000, approximately 500 MBR systems were in operation worldwide, of which 66% of commercial use MBRs were operating in Japan (Stephenson et al., 2000). The remaining membrane systems are in North America and Europe (Stephenson et al., 2000). Applications of MBR technology include treatment of municipal wastewater, process water from the food, chemical, dye, agriculture, brewery, and medical industries. Treatment objectives and performance can differ based upon sludge characteristics and discharge requirements (Brindle and Churchouse, 2001). Further, MBR systems are commercially available from a number of suppliers (e.g., Zenon, US Filter, Enviroquip, Mitsubishi) that utilize flat plate, hollow fiber, or tubular membrane technologies (Stephenson et al., 2000).

Past studies employing MBR systems indicate clear reduction of key wastewater parameters. Viadero and Noblet (2002), applying a laboratory scale membrane filter with a 0.05 mm pore size, but with no biological treatment component, saw removal efficiency of total suspended solids (TSS) of 94% and COD of 76%. Babcock et al. (2004) found that in a side-by-side analysis of four different types of pilot-scale membrane bioreactor technologies, inlet TSS levels of up to 400 mg/L were reduced to <4 mg/L. Additionally, Biological Oxygen Demand (cBOD5) removal efficiency was consistently about 99%. Removal of total nitrogen (TN) was 60-76% and total phosphorus (TP) removal was in the range of 70-85%. Similarly, in a large-scale membrane bioreactor system in Porlock, UK, Churchouse and Brindle (2003) showed comparable removal efficiencies of TSS and cBOD5. In addition, these researchers showed the capacity of MBR technology to perform bacterial and viral removal with a greater than six log reduction in bacteria and three to five log reduction in viruses reported. In a comparative analysis of both a CAS system and a MBR, the CAS

system indicated a peak TN removal efficiency of 62%, while the MBR showed a peak TN removal of 77% (Soriano et al., 2003). CAS peak COD removal was 85% while MBR COD removal was 96% (Soriano et al., 2003). A key advantage of the MBR over the CAS is the ability of the membrane to retain bacteria, which prevents the entrainment of nitrifiers/denitrifiers in the effluent (Soriano et al., 2003). Further, while the CAS requires a biosolids concentration of approximately 0.5% to prevent concentrated floc settling problems, the MBR can operate at solids concentrations of 2-3% (Marrot et al., 2004). As a result, the potential for MBRs to perform wastewater treatment at a finer scale than traditional wastewater treatment systems is clear. In scenarios with the need of a water system with the capacity to reduce key water quality parameters below stringent threshold levels or for wastewater reclamation, MBR technology appears to have possible widespread applications.

1.4. Effects of increased salinity on wastewater treatment

One particularly challenging aspect of wastewater treatment is the management of a high salinity effluent. Specifically, nitrogen compounds may accumulate because of the potential for inhibition of nitrifying and denitrifying bacteria (Sakairi et al., 1996). Diverging conclusions have been reported relating to the impact of high salinity on the activated sludge process (Hamoda and Al-Attar, 1995). In a study by Sanchez et al. (2004), where concentrations from 0 g/L to 60 g/ L NaCl were utilized, a linear decrease was reported in the rates of both nitritation (NH3! NO2~) and nitratation (NO2~ ! NO3~) with increased salinity. And, Sakairi et al. (1996) reported nitrification rates approximately six times less at higher salinity compared to freshwater. In contrast, Hamoda and Al-Attar (1995) reported no deterioration in the activated sludge process with NaCl concentrations of 30 g/L. In a similar study, Dahl et al. (1997) reported that maximum nitrification rates were achieved at 20 g/L chloride.

Similar variations associated with the effects of high salinity on denitrification have been reported. In an experiment conducted by Yang et al. (1995), utilizing an up-flow reactor to enhance denitrifying bacterial growth, nitrate removal at NaCl concentrations of 0 g/L, 10 g/L, 15 g/L, 20 g/L, 25 g/L, and 30 g/L were tested. Results indicated that denitrification capacity (%) was reduced to 75% at 20 g/L NaCl and 60% at 30 g/L NaCl when compared to the 0 g/L salinity

control. In a similar study using bench-scale sequencing batch reactors, the specific nitrate reduction rate decreased proportionally with the increase in salinity (Glass and Silverstein, 1999). Conversely, Sakairi et al. (1996) detected 100% nitrate removal under seawater conditions provided that sufficient phosphorus was available for adenosine tri-phosphate (ATP) generation.

Although little information is available relative to the impact of increased salinity on phosphorus removal, Barak and van Rijn (2000) postulated that because the primary mechanism for phosphorus removal is associated with denitrifying bacteria, similar salinity effects are likely to be observed. With regard to membrane exclusion of solids (TSS, bacteria, etc.), which is a physical screening process, increased salinity is unlikely to impact their removal. However, this should be researched to determine if changes in salt concentrations create unforeseen changes in precipitates or release of cellular by-products that could hinder permeate flow through the membrane.

1.5. Objective

The objective of this study was to evaluate the performance of a pilot-plant MBR at treating fish culture biosolids discharged from an aquaculture facility and to assess the potential for the return of processed water for reuse in the fish culture system. Salinity levels within the MBR system were manipulated to determine the effects of salinity on membrane filter function. The hypothesis to be tested: increasing salinity from <0.03 ppt to 32 ppt will have no effect on MBR performance once the system has been given sufficient time to re-acclimate to the new conditions. Specifically, analysis of outlet concentrations and removal efficiencies of the key water quality parameters will indicate no reduction in their removal at higher salt concentrations.

2. Materials and methods

2.1. Waste water source

The Membrane Filtration study was conducted at the Conservation Fund's Freshwater Institute (Shep-herdstown, West Virginia) utilizing the waste stream emanating from two recirculating aquaculture systems with a total of 35 mtonnes (80,000 lbs) of annual rainbow trout (Oncorhynchus mykiss) production (Fig. 1). The first was a partial reuse system that recirculates 1200-1850 lpm (320-490 gpm) of water through three 3.66 m (12 ft) x 1.1 m (3.5 ft) circular "Cornell-type" dual drain culture tanks, which recycled 85-90% of the total flow (Summerfelt et al., 2004). The recycled flow was collected and filtered through a rotating drum filter (Model RFM 3236, PRA Manufacturing Ltd., Nanaimo, British Colombia, Canada) equipped with 90 mm filter screens. The second wastewater source originated from a fully recirculating fish grow out system that contained a single 9.1 m (30 ft) x 2.4 m (8 ft) tank that recycled approximately 4800 lpm (1250 gpm) of water (Davidson and Summerfelt, 2005). The entire water flow through the system was collected and filtered by a rotating drum filter (Model RFM 4848, PRA Manufacturing Ltd., Nanaimo, British Colombia, Canada) equipped with 90 mm filter screens. Backwash effluent from both rotating drum filters drained into a below ground equalization tank located external to the fish culture facility (Fig. 1). Process water fed into the MBR system via the equalization tank was controlled by a pump and float switch system. When the water level in the equalization tank reached a specified depth, a float switch activated a pump, which then fed wastewater into the MBR system (Fig. 1). To achieve the desired flow through the MBR, any excess wastewater flow pumped from the equalization tank was diverted to an off-line settling basin.

Fig. 1. Schematic indicates the flow path of drum filter backwash flows from fish culture systems to the membrane biological reactor (MBR).

aerated membrane tank

Fig. 2. Drawing indicates location and orientation of the main components of the MBR system.

2.2. Membrane biological reactor system

The MBR system (Enviroquip, Austin, TX, USA) tested (Fig. 2) contained two reactor tanks; one that was maintained in an anoxic state while the other was aerobic. The design of the MBR system was generally based upon the modified Ludzack-Ettinger single sludge process (EPA, 1993a). However, the clarifier unit used in the Ludzack-Ettinger design is replaced in this process by a membrane filter submerged in the mixed liquor. The anoxic tank, dimensions 2.6 m (8.5 ft diameter) x 2.4 m (8 ft tall), provided 6760 L (1790 gal) of operating capacity and received the flow from the equalization tank. The aerobic tank, dimensions 1.5 m (5 ft diameter) x 3.0 m (10 ft tall), provided 5050 L (1340 gal)

Fig. 3. Parallel orientation of the membrane plates and tubing directing flow of processed water through permeate manifold.

of operating capacity and contained the submerged membrane unit (Kubota Manufacturing, Japan), which is capable of extracting 22.6 m3/day (6000 gal/day) of permeate from the mixed liquor solution. The rack of 50 Kubota plate membranes provided a total membrane surface area of 40 m2 (Fig. 3). Overall flux through the membrane rack was set at <0.57 m3/day m2 surface area. The membranes provided a 0.4 mm nominal pore size, which becomes even finer as biofilm coats the membrane. A Goulds (Seneca Falls, NY) 1/3 hp pump recycled approximately 54.5 m3/day of the mixed liquor from the anoxic tank to the aerobic tank. Overflow from the aerobic tank gravity fed into the anoxic tank to complete the water recirculation loop. Aeration was provided by a five horsepower Model-11 Dresser Roots blower

Fig. 4. Rolling action of the MLSS in the aerobic tank illustrates the continuous air scouring of membranes provided by course bubble aeration.

(Turnbridge, Huddersfield, England). Aeration rate below the membranes was never allowed to drop below 5.5 m3/min in order to provide continuous bubble scouring of the membranes (Fig. 4). Dissolved oxygen concentration was continuously monitored in the aerobic tank using a Danfoss Evita Oxy dissolved oxygen meter (Loveland, CO). A Proportional Integral Derivative (PID) control of blower speed was provided by an Allen Bradley SLC 500 programmable controller (Milwaukee, WI). Aeration rate was adjusted with a PID controller to maintain a dissolved oxygen concentration of approximately 2.0 mg/L. The anoxic tank was not aerated so as to maintain dissolved oxygen concentrations of less than 0.5 mg/L. Concentration of mixed liquor suspended solids (MLSS) within the anoxic and aerobic tanks was maintained at approximately 18,000-30,000 mg/L by periodic (approximately bi-weekly) biosolids removal. Permeate water was pulled through the submerged membrane unit by a Webtrol centrifugal pump (Weber Industries, St. Louis, MO). The membrane was operated 24 h daily with a repeat cycle of 9 min of permeate flow followed by 1 min of relaxation in order to maintain a relatively low trans-membrane pressure differential. An automated 20-min air-scouring event at an aeration rate of 12-13 m3/min was programmed to occur nightly to reduce build up of excess biofilm on the membranes.

2.3. Sampling regime

Water samples were taken from three sampling ports in the MBR system (Fig. 5). The first was located at the inlet into the anoxic tank from the equalization tank and was used to evaluate the characteristics of the incoming wastewater. The second sampling site was from the overflow pipe connecting the anoxic and the aerobic tanks. This site was sampled primarily for suspended solids in order to maintain a desired mixed liquor

volatile suspended solids (MLVSS) concentration. The third sampling site was located after the submerged membrane unit in the effluent permeate line. This was done in order to compare water quality characteristics of the effluent to the influent water.

Salinity levels within the membrane biological reactor system were manipulated by adding salt (NaCl) into the anoxic tank. Specifically, a Meyers Mini Salt Spreader (Cleveland, OH) mounted above the anoxic tank added a Mix-n-Fine (Cargill Salt, Minneapolis, MN) salt into the system via a timer control mechanism, which allowed for hourly addition of salt. Salinity levels in both the anoxic tank and MBR effluent were monitored daily (recorded in parts per thousand) with a YSI (Yellow Springs, OH) Model 30 Handheld Salinity, Conductivity, and Temperature System to ensure that the correct salinity was maintained. Salinity levels that were investigated were approximately 0 ppt, 8 ppt, 16 ppt, and 32 ppt. MBR operation began in May 2004 and was managed under freshwater conditions at a Hydraulic Loading Rate (HLR) of 13.6 m3/day until study initialization. The experiment was conducted from 26 October 2004 to 22 June 2005 (239 days) at a HLR of 6.8 m3/day. Ten sets of data points at each level of salinity were collected, once treatment across the MBR had reached quasi-steady-state conditions. Time periods for data collection once quasi-steady-state conditions were reached at each treatment were days 225-261, 267-303, 420-442, and 448-464 for 0 ppt, 8 ppt, 16 ppt, and 32 ppt salinity, respectively.

2.4. Water quality parameters analyzed

The three sampling sites were tested for a series of water quality parameters (Table 1). Methods were assessed based upon salinity interference. Seawater is indicated as a source of interference when applying

Fig. 5. Schematic indicates sampling ports and the flow of wastewater within the membrane biological reactor system.

Table 1

Laboratory methods used for each water quality parameter (APHA, 1998), units expressed, and sampling locations

Parameter Method Units Sampling location

Salinity YSI Model 30 Handheld ppt 1-3

Dissolved oxygen (DO) Hach Model HQ10LD0 mg/L 1-3

pH YSI Model 60 Handheld pH Meter pH 1-3

Alkalinity Standard Methods 2320 mg/L (as CaC03) 1, 3

Total nitrogena b c Calculated mg/L 1, 3

Total ammonia nitrogenb,c,d,e Standard Methods 4500-NH3 mg/L (as NH3-N) 1, 3

Nitrogen-nitrited,e Standard Methods 4500-N02 mg/L (as N02-N) 1, 3

Nitrogen-nitrateb,d,e Standard Methods 4500-N03 mg/L (as N03-N) 1, 3

Organic nitrogena,b Calculated mg/L 1, 3

Total kjeldahl nitrogenb Standard Methods 4500-Norg mg/L (as TKN-N) 1, 3

Total phosphorusb,c,e Standard Methods 4500-P mg/L 1, 3

Total suspended solidsc,d Standard Methods 2560 mg/L 1-3

Total volatile solidsb,c Standard Methods 2560 mg/L 1-3

cB0D5b Standard Methods 5210 5-day B0D mg/L 1, 3

Total coliformb Hach membrane filtration method 8074 cfu/100 mL 1, 3

Total heterotrophsb Hach membrane filtration method 8242 cfu/mL 1, 3

a Calculated based upon values obtained for total kjeldahl nitrogen, total ammonia nitrogen, nitrite, and nitrate.

Removal efficiency calculated. c Analysis of variance performed.

d Standard additions performed to assess error associated with salinity e Analyzed with a DR4000/U spectrophotometer.

Standard Methods 4500-NH3 for total ammonia nitrogen (TAN) (APHA, 1998). To calculate error, standard additions were performed on the effluent samples at the higher salinities, indicating 75% recovery at 32 ppt salinity. As a result, at the higher salinities, reported effluent TAN concentrations are potentially low by 25%. Seawater is also indicated as a source of interference when applying Standard Methods 4500-N03 to assess nitrate nitrogen (APHA, 1998). Standard additions were performed on the effluent samples at the higher salinities to calculate error, which indicated 50% recovery at 32 ppt salinity. Consequently, reported high salinity effluent nitrate-nitrogen concentrations are potentially low by 50%. The Hach HQ10 LDO meter used to measure dissolved oxygen in the test were compensated for salinity. Enumeration of heterotrophic and total coliform bacteria was conducted at sampling sites #1 (inlet) and #3 (effluent). During each sampling event, two or three replicates were assayed for total heterotrophic bacteria and total Coliform bacteria at both sampling sites. Heterotrophic bacteria were assessed utilizing Hach membrane filtration method 8242 using m-TGE Broth with TTC indicator. After incubation, colonies were counted with a low-power microscope and reported in number of colony forming units (cfu) per 1 mL sample. Similarly, coliform bacteria were analyzed using Hach Membrane Filtration method 8074 (m-Endo Broth). Colonies were

counted with a low-power microscope and reported in cfu per 100 mL sample. No indication of interference is attributed to high salinity when applying either bacteria enumeration method (APHA, 1998).

Data were collected and compiled for assessment based upon the treatment efficiency of the MBR at each of the salinity levels. Each of the water quality parameters are expressed in terms of their mean ± standard error. Removal efficiencies of each key water quality parameter are calculated (i.e., ((inlet — outlet)/inlet) x 100) and compared based upon salinity level (Table 1). An analysis of variance (ANOVA) was conducted separately for the most interesting quality parameters (Table 1) in order to determine statistical differences in the mean effluent concentrations at each salinity level. Specifically, four mean outlet concentrations were calculated representing each of the salinity concentrations (e.g., TSS mean in mg/L at 0 ppt, 8 ppt, 16 ppt, 32 ppt) and analyzed for differences in the means.

2.5. Activated sludge process assessment

The mean cell residence time (Uc) or sludge age and the food to microorganism ratio (F:M) are two common parameters that can provide insight into the design and control of an activated sludge process (Metcalf and Eddy, 1991). A high mean cell residence time and a low

F:M will produce a lower sludge yield (Stephenson et al., 2000). Mean cell residence time for the MBR system was calculated (Eq. (7)) based upon Stephenson et al. (2000) as follows

Qw Xw + QeXe

where Uc is the mean cell residence within the MBR system (days), Vr the MBR system volume (m3/day), X the concentration of volatile suspended solids in the MBR system (mg/l), Qw the waste sludge removed (kg/ day), Xw the concentration of volatile suspended solids in the waste sludge (mg/l), Qe the treated effluent flowrate (m3/day) and Xe is the concentration of volatile suspended solids in the treated effluent (mg/l).

The food to microorganism ratio was calculated according to Metcalf and Eddy (1991) (Eq. (8)) as follows

F : M =

where F:M is the food to microorganism ratio (day- ), S0 the inlet cBOD5 (mg/l), U the hydraulic detention time based on the MBR system volume = Vr/Qe (days), and X is the concentration of volatile suspended solids in the MBR system (mg/l).

3. Results and discussion

3.1. MBR operation experience

We found that a key advantage to the MBR system was its relative ease of operation and lack of extensive maintenance requirement. The automated monitoring features allow for minimal personnel commitment. Specifically, dissolved oxygen requirements were maintained under optimum conditions over months-long time periods by the dissolved oxygen monitor and the proportional integral derivative (PID) controller. Moreover, float switches in the anoxic tank allow the MBR to maintain proper depth, processing permeate water and "requesting" drum filter backwash flows from the equalization tank as needed. Membrane fouling is automatically mitigated through programmable logic controller (PLC) procedures in which daily membrane air scouring events prevent excessive build up of biological material. Further automation of optimized permeate flux through the membranes involves the ability to program permeate pump run/ relax cycling. A 9 min run followed by a 1 min relax cycling of the permeate pump allows flux of processed water through the membranes for 9 min with relaxation

Fig. 6. Trans-membrane pressure (TMP) over the course of the study and chemical cleaning events with (1) sodium hypochlorite and (2) HCl. Membrane flux was 0.2 lpm/m2 membrane surface area.

and air scouring for 1 min. This automated process sustains membrane flux over extended periods with little operator involvement (Fig. 6).

Operator maintenance duties were minimal. Daily checks were required to ensure proper function of critical components (pumps, mixer, and blower unit), verify manufacturer's recommended trans-membrane pressure range, and confirm dissolved oxygen levels in both the aerobic and anoxic tanks. Approximately weekly solids removal events from the MBR system were performed to maintain MLSS within the desired range. In particular, this 15-min procedure involved diverting recycle pump flow from the aerobic tank into a settling cone for later land application. Uninterrupted MBR use was maintained throughout the solids removal procedure. Bi-annual chemical membrane cleaning to reduce biofouling and CaCO3 precipitation is recommended by the manufacturer and was confirmed by experience (Fig. 6). Membrane fouling was monitored by periodically recording the trans-membrane pressure (TMP) value at the end of a 9-min permeate pumping cycle. The TMP is actually a vacuum pressure that is produced as the permeate pump suctions water out of the membrane. A mean TMP of 1.4 ± 0.1 psi was observed over the course of the experiment. Fig. 6 illustrates TMP trend over the course of the experiment and indicates membrane chemical cleaning events. In situ chemical cleaning was simultaneously performed on all membrane cartridges with a solution gravity fed to the membranes from an external tank. A 189 L (50 gal) 0.5% sodium hypochlorite solution was used to reduce biofouling on two separate occasions, while a single 189 L (50 gal) 5% hydrogen chloride solution was used to dissolve inorganic scaling. A 1-2 h interruption of MBR operation was necessary to perform chemical cleaning. Further, no negative effect

on microbiological removal capacity was observed subsequent to chemical cleaning of the membranes, confirming claims of the membrane system supplier. Membrane cleaning procedures were found to be simple and effective.

During the experiment, solids removal events were performed every 6.8 ± 0.7 days with a mean volume removed of 1.49 ± 0.1m3. Mean concentration of MLVSS in the sludge removed was 18,857 ± 628 mg/L and a mass MLVSS removed of 27.7 ± 1.9 kg/event, resulting in a rate MLVSS removed of 7.5 ± 1.6 kg/day. Using Eq. (7), a mean solids detention time (Uc) of 64 ± 8.0 days was calculated. Because the mass of TVSS flushed out of the MBR within the permeate flow was negligible (i.e., 0.1 kg/day) relative to the mass of TVSS retained by the membrane (i.e., 27.7 ± 1.9 kg), the product of treated effluent flow rate times the concentration of volatile suspended solids in the treated effluent (QeXe) was negligible and assumed to be zero. Using Eq. (8), the mean F:M ratio was calculated as 0.029 day—1. Typical waste treatment plants that process municipal wastewater and utilize a CAS system have a mean cell residence time of3-15 days and a F:M ratio of 0.05-1.0 day— 1 (Metcalf and Eddy, 1991). MBR technology has the ability to operate at mean cell residence time of 6.2 days to >100 days and F:M ratios in the range of 0.05-0.15 day—1 (Stephenson et al., 2000). Comparing the F:M ratio used in the present study to that recommended by others indicates that the MBR could have been loaded with two to six times more cB0D5 and would have remained within acceptable F:M ratio.

0ver the course of the experiment, mean dissolved oxygen concentrations (DO) were 3.2 ± 0.3 mg/L in the aerobic tank and 0.11 ± 0.02 mg/L in the anoxic tank. In our experience, the MBR works best when fully loaded with all waste solids coming from the drum filter. This is because the membranes require a minimum

aeration rate below the membranes to scour them clean and lower cBOD5 loading rate reduces oxygen demand. Ideally, the MBR is operated to maintain a DO concentration of 2 mg/L in the aerobic membrane tank and a DO of near 0 mg/L in the anoxic tank. If cBOD5 loading on the MBR is too low, then this minimum aeration rate is higher than is required for cBOD5 removal and the dissolved oxygen concentration increases to 4-6 mg/L in the aerobic tank. This can create a problem when the oxygenated water is recirculated back to the anoxic tank, because it will raise the DO in the anoxic tank and the higher DO can reduce denitrification. In addition, the MBR is operated at a mixed liquor volatile suspended solids (MLVSS) concentration of 15,00030,000 mg/L. So the membranes are always seeing a high solids loading. Therefore, pre-treating the backwash flow does not make sense, because the inlet TSS is only about 1000 mg/L, which is much lower than the MLSS around the membranes.

Mean alkalinity in the MBR was 275 ± 5 mg/L in the inlet and 305 ± 5 mg/L in the permeate, indicating recovery of alkalinity across the waste treatment system. Theoretical stoichiometry indicates that for every 1g of NH4+-N consumed by nitrifying bacteria 7.1 g alkalinity (as CaCO3) are destroyed, and for every 1 g NO3—-N consumed by denitrifiers 3.57 g alkalinity (as CaCO3) are produced (EPA, 1993a). There was little nitrate entering the MBR, but a NO3—-N concentration of only 10 mg/L would have explained the net production of alkalinity measured across the MBR system. We speculate that the array of micro-biological pathways that were involved in the conversion of the waste protein to TAN to cell mass to nitrite or nitrate may have accounted for this net increase in alkalinity across the MBR.

The MBR system footprint (153 m2), including working room around the equipment, was small relative to the fish culture facility footprint (1829 m2), resulting in an 8.4% space requirement for MBR treatment of

Table 2

TSS and TVS removal at all conditions

Salinity (ppt)

Inlet (mg/L) ± S.E. Outlet (mg/L) ± S.E. Removal (%)

Inlet (mg/L) ± S.E. Outlet (mg/L) ± S.E. Removal (%)

1688 ± 302 0.3 ± 0.1

99.98 ± 0.01

1380 ± 246 0.1 ± 0.04

99.99 0.0

1732 ± 436 1.2 ± 0.2 99.90 ± 0.2

1454 ± 357 0.4 ± 0.1 99.96 ± 0.0

1357 ± 296 1.3 ± 0.1 99.83 ± 0.03

1144 ± 257 0.2 ± 0.04 99.97 ± 0.01

754 ± 64 2.5 ± 0.7 99.65 ± 0.1

642 ± 55 0.3 ± 0.1 99.96 ± 0.01

biosolids compared to total area for fish culture. Further, the ability to site the MBR in a location removed from the fish culture facility allows for biosolids treatment and water reclamation in a biosecure setting.

3.2. Total suspended solids

Operated at a MLSS concentration of 15,00030,000 mg/L over the course of the experiment, the MBR showed highly efficient removal (>99%) of TSS and TVS at all salinity levels (Table 2). Even visual inspection of water quality showed profound differences (Fig. 7). Mean outlet concentrations of TSS at 0 ppt, 8 ppt, 16 ppt, and 32 ppt were 0.3 ± 0.1 mg/L, 1.2 ± 0.2 mg/L, 1.3 ± 0.1 mg/L, and 2.5 ± 0.7 mg/L, respectively. And, outlet TVS concentrations were 0.1 ± 0.04 mg/L, 0.4 ± 0.1 mg/L, 0.2 ± 0.04 mg/L, and 0.3 ± 0.1 mg/L, respectively. An analysis of variance (ANOVA) conducted across the salinities indicated significant difference (p < 0.001, a = 0.05) in mean TSS outlet concentrations. However, post hoc analysis utilizing Fisher's PLSD showed no difference (p = 0.4261, a = 0.05) in mean outlet TSS concentration when comparing means at 8 ppt and 16 ppt salinity.

A slightly reduced capacity for solids removal was statistically significant at increased salinity. A possible explanation for increased TSS concentrations in permeate flow may be that elevated salinity can have a negative effect on membrane integrity and solids removal potential. However, further research with regard to this mechanism needs to be conducted. Alternatively, this occurrence may be related to ordinary membrane deterioration over the course of the experiment.

3.3. Bacteria removal

The membrane biological reactor was also efficient at exclusion of bacteria from permeate flow at all treatment levels (Table 3). Treatment efficiencies of total heterotrophs ranged from 2 to 5.6 log10 removal

over the course of the experiment. Results indicate total heterotroph bacteria counts in the permeate ranged from 2 cfu/mL to 121 cfu/mL. Removal efficiencies of total coliform ranged from 3.2 to 7.0 log10 removal at the four treatment levels. Further, total coliform bacteria counts in the permeate ranged from 0 cfu/mL to 80 cfu/ mL over the course of the experiment. Bacteria in the permeate could have been either the result of biofilm re-growth in the permeate piping or a hole or tear in the membrane.

Further investigation of the bacteria enumeration procedures for total heterotrophs and total coliform and the potential error associated with analysis under seawater conditions indicate that results are precise relative to freshwater samples. According to Standard Methods—Method 9222 (APHA, 1998), the membrane filter (MF) technique is highly reproducible and useful for monitoring bacteria counts in natural water systems, including saline water. Standard Methods (APHA, 1998) does refer to utilization of a buffered solution when analyzing samples containing high concentrations of heavy metals. However, this is not relevant with regard to water samples high in Na+ ions.

3.4. Biochemical oxygen demand removal

Results indicate that the MBR system was highly efficient at removing cBOD5 under all of the conditions tested (Table 4). Outlet concentrations were consistently low during all salinity trials. Specifically, mean cBOD5 outlet concentrations ranged from 0.6 mg/L to 1.3 mg/L for all the trials. Mean cBOD5 removal exceeded 99.8% at all salinity levels (Table 4). As a result, evidence suggests that increased salinity had no effect on the cBOD5 removal capability of the heterotrophic microorganisms in the MLSS. A coupled cBOD5 removal/denitrification process is evident in the anoxic/aerobic sequence of the MBR system. Facultative denitrifiers utilize NO3~ to metabolize exogenous carbon (as influent cBOD5), and sufficient C/N ratios are necessary to drive this process (EPA, 1993a).

Table 3

Removal of total heterotrophs and total coliform in the MBR system

Salinity (ppt)

Total heterotrophs Inlet (cfu/mL) ± S.E. Outlet (cfu/mL) ± S.E. Removal (%)

Total coliform

Inlet (cfu/100 mL) ± S.E. Outlet (cfu/100 mL) ± S.E. Removal (%)

3.4E+6 ± 1.3E+6 68 ± 26 99.993 ± 0.004

1.4E+7 ± 5.2E+6 1 ± 1 99.996 ± 0.004

1.2E+6 ± 5.3E+5 121 ± 70 99.95 ± 0.02

1.0E+7 ± 4.9E+6 80 ± 38 99.94 ± 0.05

1.5E+6 ± 5.8E+5 8 ± 1 99.997 ± 0.001

1.3E+7 ± 5.5E+6

1 ± 0.4 9.99999 ± 0.00001

1.7E+6 ± 9.1E+5 2 ± 1 99.9998 ± 0.0001

5.7E+6 ± 1.4E+6 0 ± 0.2 99.9998 ± 0.0001

Research indicates that cBOD5/TKN ratios of 8.7 and 11.9 correspond to nitrogen removals rates of 66% and 83%, respectively (EPA, 1993a). The high inlet cBOD5 from drum filter backwash flows in this experiment (Table 4) resulted in a cBOD5/TKN ratio of approximately 15 and facilitated the coupled cBOD5 removal/denitrification process. We observed TN removal rates of 89.5-95.5% (Table 5), indicating comparable removal efficiency across the salinity treatments.

3.5. Nitrogen removal

The MBR system performed similarly with regard to the removal of nitrogen under all conditions tested (Table 5), provided that a sufficient acclimation period to the increased salinity was allowed. Although TAN removal percentages varied based upon inlet concentrations, mean outlet concentrations were consistently low. Specifically, mean outlet TAN concentrations were 1.4 ± 0.7 mg/L, 1.8 ± 0.3 mg/L, 0.9 ± 0.1 mg/L, and

Table 4

Efficiency of cBOD5 removal by the MBR at all conditions

Salinity (ppt)

Inlet (mg/L) ± S.E. Outlet (mg/L) ± S.E. Removal (%)

1075 ± 145 1 ± 0.9 99.996 ± 0.004

1583 ± 410 1.3 ± 0.03 99.91 ± 0.02

930 ± 281 0.7 ± 0.1 9.87 0.04

372 0.6 99.84

Table 5

Removal of TAN, TN, and organic nitrogen in the MBR system

Salinity (ppt)

0 8 16 32

Inlet (mg/L) ± S.E. 4.1 ± 0.7 1.9 ± 0.3 4.5 ± 2.5 2.6 ± 0.4

Outlet (mg/L) ± S.E. 1.4 ± 0.7 1.8 ± 0.6 0.4 ± 0.1 2 ± 1.2

Removal (%) 57.4 ± 19.2 13.7 ± 22.8 78.3 ± 6.6 28.1 ± 42.0

Total nitrogen

Inlet (mg/L) ± S.E. 68.4 ± 12.9 67.3 ± 3.6 62.4 ± 22.4 50.7 ± 5.7

Outlet (mg/L) ± S.E. 3.9 ± 1.3 3.1 ± 0.7 2.6 ± 0.4 2.0 ± 0.2

Removal (%) 91.8 ± 2.9 93.0 ± 2.0 93.9 ± 1.0 95.5 ± 0.6

Organic nitrogen

Inlet (mg/L) ± S.E. 63.4 ± 12.6 63.4 ± 14.8 57.3 ± 19.9 46.0 ± 5.4

Outlet (mg/L) ± S.E. 1.8 ± 0.8 0.9 ± 0.1 1.3 ± 0.2 0.01 ± 0.01

Removal (%) 95.6 ± 1.9 97.9 ± 0.5 96.7 ± 0.7 99.98 ± 0.02

2.0 ± 1.2 mg/L at 0ppt, 8 ppt, 16 ppt, and 32 ppt, respectively. Analysis of variance (ANOVA) indicates that there was no significant difference between the four mean TAN outlet concentrations ( p = 0.5437, a = 0.05). Additionally, results indicate that the MBR system was efficient at removing total nitrogen, which was accomplished without supplemental carbon substrate addition. Analysis of the TN outlet concentrations also indicates efficient removal. In particular, mean outlet TN concentrations were 3.9 ± 1.3 mg/L, 3.6 ± 0.7 mg/L, 3.5 ± 0.7 mg/L, and 2.0 ± 0.2 mg/L at 0 ppt, 8 ppt, 16 ppt, and 32 ppt, respectively. Analysis of variance (ANOVA) indicates that there was no significant difference between the four mean TN outlet concentrations ( p = 0.5855, a = 0.05) at the four treatment levels. The removal percentages of organic nitrogen were 95.6 ± 1.9, 97.9 ± 0.5, 96.7 ± 0.7, and 99.98 ± 0.02 at 0ppt, 8 ppt, 16 ppt, and 32 ppt, respectively.

A microbiological turnover period was observed as nitrifying bacteria (NH3 ! NO3) acclimated to salinity increasing from 8 ppt to 16 ppt. A prolonged bacterial acclimation period under freshwater conditions was conducted prior to collecting data at 0 ppt salinity. When salinity within the MBR was increased to 8 ppt, no reduction in nitrogen removal capacity was observed (Fig. 8), indicating little to no effect on nitrifying bacteria. However, when salinity was increased from 8 ppt to 16 ppt, a 110-day acclimation period to the new condition was necessary before steady-state nitrification

was achieved. Fig. 8 indicates changes in permeate TAN, nitrate-nitrogen, and nitrite-nitrogen concentrations over the course of the experiment. Data indicate that ammonia-oxidizing bacteria are vulnerable to increased concentration in salinity beyond 8 ppt. Specifically, results suggest that the transition from 8 ppt to 16 ppt salinity causes a decrease in the Nitrosomonas spp. population. Similar observations were made by Chen et al. (2003) when analyzing nitrifier response to increased salinity. Chen et al. (2003) reported that specific nitrification rate (mg-N/g MLVSS/h) decreased when chloride concentration was increased from 10,000 mg chloride/L to 20,000 mg chloride/L. A 4-week acclimation period was needed for saline adapted ammonia oxidizers to build up. Hovanec and DeLong (1996) suggest that with regard to Nitrosomonas spp., ammonia-oxidizing bacteria in freshwater aquaria are a different species from ammonia-oxidizing bacteria in seawater aquaria. Therefore, it is likely another species of ammonia-oxidizing bacteria was cultivated in the membrane biological reactor before complete nitrification could recover. Additionally, with reference to Nitrobacter spp., several subdivisions of Proteobacteria are capable of nitrite oxidation (Hovanec and DeLong, 1996). As a result, a possible account for uninterrupted nitrite uptake is explained by evidence of a consortium of bacteria proficient at nitrite oxidation at a wide range of salinities. Alternatively, continued NO2~ uptake at increasing salinity may have been (at least in part) due to direct conversion to N2 by denitrifying heterotrophic

0 10 20 30 40 50 60 70 80 90 100 110 Day

Fig. 9. TAN, NO3~-N, NO2~-N concentrations during MBR start up.

bacteria. In either scenario, bacteria capable of withstanding changes in salinity may have performed further, compensatory nitrite uptake.

The significant increase in permeate TAN concentration observed during the transition from the 8 ppt salinity trial to the 16 ppt salinity trial was initially considered to be similar to the transient spike in TAN that can occur during startup of a nitrifying bacteria population within a more traditional biofilter, i.e., where an initial spike in the concentration of TAN is followed by a subsequent peak in nitrite and then by an increase in nitrate. However, the transition from the 8 ppt salinity trial to the 16 ppt salinity trial did not produce an appreciable increase in nitrite or nitrate concentrations. To explain these results, we hypothesize that the heterotrophic bacteria responsible for denitrification were not inhibited by the increase in salinity, whereas the autotrophic nitrifying bacteria were inhibited, at least to some extent. In comparison, when the MBR was first started-up, the TAN, nitrite, and nitrate concentrations never conclusively spiked—outside of one data set that was collected some 60 days later (Fig. 9). We think that the rapidly growing population of heterotrophic bacteria incorporated a large portion of the TAN that was released when the protein in the waste solids degraded within the MBR. Further, in order to reduce foaming during start up, sugar was introduced as a carbon source to facilitate bacterial growth. As a result,

it is possible that increased TAN utilization was coupled with carbon amendment.

Volatilization of gaseous NH3 due to vigorous aeration in the aerobic tank is not considered a significant mechanism for ammonia removal under the conditions tested. Temperature conditions in the MLSS ranged from 16.6 °C to 22.3 °C while pH ranged from 7.13 to 7.48 over the course of the experiment. And, according to Piper et al. (1982), these conditions would result in a percent of unionized ammonia (NH3) in the mixed liquor ranging from 0.37% to 1.43%, which would not allow for significant stripping of NH3.

3.6. Phosphorus removal

A marked reduction in phosphorus removal by the MBR system was observed at increased salinity (Table 6). The outlet concentration of phosphorus at 0 ppt salinity was 1.5 ± 0.3 mg/L. However, outlet concentrations at 8 ppt, 16 ppt, and 32 ppt salinity were 8.2 ± 0.5 mg/L, 6.2 ± 0.3 mg/L, and 6.2 ± 0.5 mg/L, respectively. Analysis of variance (ANOVA) indicates a significant difference exists between mean outlet concentrations ( p < 0.0001, a = 0.05). Data suggest that elevated salinity inhibited luxury phosphorus uptake. The enhanced biological phosphorus removal (EBPR) process by de-nitrifying bacteria in the activated sludge was suppressed. Reduced removal at all elevated salinities was seen regardless of prolonged acclimation periods. Houghton et al. (1971) also observed inhibition of phosphorus uptake at elevated salinity. They observed dramatically reduced phosphorus removal in the activated sludge process at 10 ppt salinity when compared to 1 ppt salinity. The sodium ion (Na+) was thought to be responsible for the phenomenon, because similar results were seen when either NaCl or NaHCO3 were added to the activated sludge (Houghton et al., 1971). An alternative saline aquaculture effluent nutrient removal technology evaluated by Lymbery et al. (2006) utilizing a halophyte plant (Juncus krausii) cultivated in an artificial wetland also indicated reduced phosphorus removal capacity at

Table 6

Phosphorus removal at all conditions

Salinity (ppt)

0 8 16 32

Total phosphorus

Inlet (mg/L) ± S.E. 57.2 ± 14.4 38.7 ± 8.6 24.6 ± 4.9 19.2 ± 1.9

Outlet (mg/L) ± S.E. 1.5 ± 0.3 8.2 ± 0.5 6.2 ± 0.3 6.2 ± 0.5

Removal (%) 96.1 ± 1.0 72.7 ± 3.5 70.4 ± 2.3 65.2 ± 5.4

elevated salinity (31 ppt). The authors postulated that since adsorption to soil binding sites provided the bulk of the phosphorus removal capability, binding sites became saturated over time. Further, the halophyte plant is adapted to estuarine conditions (20 ppt salinity) and exhibits reduced growth at higher salinities (Lymbery et al., 2006).

Optimization of the biological phosphorus removal process would likely be increased with the incorporation of an anaerobic reactor, which would select for phosphorus accumulating organisms (PAOs) in an environment low in DO, NO2—, and NO3— and limit competition from heterotrophic denitrifiers for carbon substrate (Reddy, 1998; EPA, 1993a; Metcalf and Eddy, 1991; Albertson, 1983). A potential retrofit of the MBR process used for this experiment might include a pre-anoxic anaerobic reactor in which PAOs would receive the highest carbon levels in the form of inlet cBOD5. The anaerobic reactor would gravity flow into the anoxic tank. In addition, a small recirculated flow would be required to provide low DO/low NO3— MLSS from the anoxic to the anaerobic tank. However, consideration should be given to the competing requirements of combined biological nitrogen and phosphorus removal processes. Specifically, a high cBOD5:TP ratio (>20:1) will ensure sufficient substrate is available for PAOs in the anaerobic reactor while also maintaining carbon for denitrifiers in the anoxic reactor, preventing accumulation of NO3— in the anaerobic reactor (EPA, 1993a). In this experiment, the cBOD5:TP ratio was approximately 28:1, indicating sufficient carbon is available to drive both processes. Consideration should also be given for competing optimal sludge detention times (Uc) of a combined nutrient removal strategy. Efficient phosphorus removal is realized at a shorter Uc relative to nitrogen removal. As a result, an activated sludge system should be operated at the shortest Uc possible that still achieves effluent nitrogen requirements (EPA, 1993a).

3.7. Alternative aquaculture biosolids reclamation technologies

A range of technologies are currently available that have the capacity to dewater concentrated biosolids flows from fish culture systems and allow reclamation of the resulting supernatant or permeate flow, each of which have their own positive and negative attributes. Typically, the concentrated backwash from recirculating aquaculture systems is treated to dewater and concentrate the waste biosolids. A belt filter de-watering apparatus employing a coagulation/flocculation technique has the

potential to effectively remove TSS and soluble phosphorus from drum filter backwash flows while showing significant removal of total nitrogen and cBOD5 (Ebeling et al., 2006). Also, geotextile tubes (constructed of a porous, woven polyethylene material) utilizing a similar polymer addition technique indicated effective TSS, TP, and TN removal (Schwartz et al., 2004). The most common dewatering method is simply to use a settling basin that has been sized to provide an extended biosolids storage period that allows the settled solids to compact (Chen et al., 1997), but the supernatants overflowing these gravity thickening tanks are high in suspended solids and dissolved wastes and would require further treatment before discharge or reuse (Brazil and Summerfelt, 2006). Vertical flow created wetlands have also been used to effectively dewater the solids contained in the backwash discharged from pilot-scale (Summerfelt et al., 1999) and full-scale aquaculture applications, but have the same disadvantages of the previously listed technologies. All of these systems possess the advantage of low energy consumption relative to the MBR system. However, the MBR is capable of waste treatment at a much finer scale, particularly with regard to removal of bacteria and dissolved wastes, such as cBOD5, nitrate, and ammonia. Additional research has been conducted specific to denitrification in marine RAS in an effort to maximize water reuse capabilities. In a literature review by van Rijn et al. (2006), denitrification reactors operating within the recirculating loop have been implemented on a limited and experimental basis. Specifically, packed bed, moving bed, and fluidized bed denitrifying reactors utilizing a variety of substrate media and carbon sources indicate the potential for incorporation of denitrification reactors within RAS. However, nitrate removal rates in these systems are varied, which necessitates further study for proper design of denitrifying reactors (van Rijn et al., 2006).

An application of denitrification within a closed marine recirculating aquaculture system utilizing the high concentration of sulfate in seawater in combination with the dissolved organic matter released by digestion of waste biosolids to drive autotrophic assimilation of nitrogen is described by Tal and Schreier (2004). Specifically, gilthead seabream (Sparus aurata) were raised in two 4.2 m3 tanks operated with a 2.0 m3 moving bed bioreactor for nitrification with a side-loop fixed bed denitrification reactor (1.0 m3). In the side loop, biosolids were collected from the backwash of a microscreen drum filter and digested in a sludge collection unit, which fed the anaerobic denitrification reactor. Results indicated that nitrate levels in this experimental system peaked at 35-45 mg/L NO3-N

while a control fish culture system (operated without a denitrification side loop) reached a concentration of 102 mg/L NO3-N (Tal and Schreier, 2004). Adding a waste biosolids digester and denitrification reactor to a closed seawater RAS can effectively retain seawater, producing a near zero-exchange system that was shown to be effective at low to moderate fish culture densities. Although this internal sludge digestion and denitrifica-tion approach is a practical alternative to MBR technology, it does add cost, complexity, and increased footprint to each RAS. Further, the internal biosolids digestion and denitrification technologies must also be distributed to every RAS and cannot be located in a single centralized location, as is the case for the MBR. If the internal biosolids digestion and denitrification technologies were centrally located to treat the biosolids discharge from multiple RAS, the treated water would pose a biosecurity threat if the flow were returned to each of the separate RAS, because this technology does not provide microbial exclusion.

The turnkey MBR system installed at Conservation Fund's Freshwater Institute, designed to treat 6000 GPD (23 m3/day), cost approximately $80,000. An MBR system designed to treat wastewater from a commercial scale aquaculture facility (approximately 450 mtonnes/year), would have to treat approximately 260 m3/day of backwash flow for reclamation, assuming a total recirculating water flow of 87,000 m3/day and a 0.3% backwash flow (87,000 x 0.003 = 260 m3/ day). A turnkey MBR system sized to reclaim this flow would cost approximately $470,000, not including cost of enclosing the system. Annual operating cost (at $0.08 kWh-1) is estimated at $20,000 (50% for electrical, 30% for sludge hauling, and 20% for labor) (Brian Codianne, Enviroquip, Austin, TX). Indeed, this is a significant capital investment. However, the potential exists with MBR technology to treat effluent to within stringent standards and combines the opportunity of reusing saline water in a marine fish culture system.

4. Conclusions

The potential of membrane biological reactors for reclaiming saline water from the biosolids back-washed from marine recirculating aquaculture systems is apparent. The MBR performed exceptionally well during this study. The physical exclusion of TSS and bacteria (total heterotrophs and total coliform) from the MLSS was nearly complete. Further, the associated cBOD5 was almost completely removed. Biological treatment of nitrogen through nitrification/

denitrification indicated consistent removal of total nitrogen at all treatment levels, provided that sufficient acclimation to each salinity level was given. On the other hand, phosphorus removal did appear to be reduced at the three higher salinity levels. Additional research will be necessary to investigate the implications of increased salinity on phosphorus removal.

Further treatment of the reclaimed water processed by the MBR system may be necessary based upon the intended use. Temperature regulation with heat exchangers may be needed in order to adjust reused water to the appropriate temperature conditions. However, if an objective of applying this technology is the culture of warm water finfish, then the conservation of heat in the membrane biological reactor process can be viewed in terms of an economic benefit. Supplementary disinfection of reclaimed water through UV irradiation and/or ozonation prior to reuse in a fish culture system might also be appropriate in order to ensure complete sterilization of potential fish pathogens. Although fixed and variable costs for the MBR system used for this experiment were relatively high, broader application of this technology in the future will likely result in a cost reduction.


Funding for this research was provided by the Agriculture Research Service of the United States Department of Agriculture, under agreement no. 591930-1-130. We would like to thank Michael Gearheart, Susan Glenn, Christine Marshall, and Angela Crone for their assistance with water quality analyses, and Brian Mason, Daniel Coffinberger and Frederick Ford for their assistance setting up and modifying the research system.


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