Recovery from Brackish Aquaculture
Recirculation System
Cover by Xuedong Zhang and J. F. Krook Printed by Delft Academic Press
ISBN
Copyright © 2014 by Xuedong ZHANG Email: x.zhang@tudelft.nl
x.d.zhang@hotmail.com
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Recovery from Brackish Aquaculture
Recirculation System
Proefschrift
ter verkrijging van de graad van doctor aan de Technische Universiteit Delft,
op gezag van de Rector Magnificus prof. ir. K.C.A.M. Luyben, voorzitter van het College voor Promoties,
in het openbaar te verdedigen op
maandag 15 december 2014 om 10:00 uur
door Xuedong ZHANG,
Master of Science in Municipal Engineering, Harbin Institute of Technology,
Prof.dr.ir. J.B. van Lier en copromotor:
Dr.ir. H.L.F.M. Spanjers
Samenstelling promotiecommissie:
Rector Magnificus voorzitter
Prof.dr.ir. J.B. van Lier Technische Universiteit Delft, promotor Dr.ir. H.L.F.M. Spanjers Technische Universiteit Delft, copromotor Prof.dr.ir. L.C. Rietveld Technische Universiteit Delft
Prof.dr.ir. M.C.M. van Loosdrecht Technische Universiteit Delft
Prof.dr.ir. A.J. Wang Harbin Institute of Technology, China Prof.dr.ir. G. Zeeman Wageningen Universiteit
Prof.dr.ir. A. Gross Ben-Gurion University of the Negev, Israel Prof.dr.ir. W.G.J. van der Meer Technische Universiteit Delft, reservelid
Introduction and scope of the thesis ... 1
Chapter 1. Potentials and Oimitations for biomethane and phosphorus recovery from Chapter 2. sludges of brackish/marine aquaculture recirculation systems: A review... 11
Performance of inorganic coagulants in treatment of backwash waters from a Chapter 3. brackish aquaculture recirculation system and digestibility of salty sludge ... 43
Effects of salinity and FeCl3on phosphatase activity, extracellular polymeric Chapter 4. substance, soluble microbial product, and specific methanogenic activity.... 63
Improving methane production and phosphorus release in anaerobic Chapter 5. digestion of particulate saline sludge from a brackish aquaculture recirculation system ... 77
Bioenergy production and bio-community distribution in a digester treating Chapter 6. sludge from a brackish aquaculture recirculation system... 89
Struvite crystallization under brackish aquaculture condition ... 119
Chapter 7. Summary and general discussion ... 135
Chapter 8. +RRIGVWXN 9. Nederlandse samenvatting en algemene discussie ... 143
List of Nomenclature... 151 List of Abbreviations ... 153 Supplementary material... 155 Acknowledgment ... 157 List of publications ... 161 Biography...... 163
1.1 Brackish/marine aquaculture recirculation system and anaerobic
digestion
Recirculation aquaculture systems (RAS) are considered a sustainable production method to meet future demand for seafood. Generally, three types of RAS configurations are employed and/or researched, as shown in Figure 1.1: A) conventional RAS without a denitrification unit; B) RAS with a denitrification unit fed with an external carbon source; C), a currently proposed RAS with a denitrification unit making using of volatile fatty acids volatile fatty acids from anaerobic digestion (AD) of solids separated from the mechanical filtration units (Aboutboul et al., 1995; Menasveta et al., 2001; Skjølstrup et al., 1998; Timmons & Ebeling, 2007). A typical RAS usually contains separation units such as, a drumfilter to intercept suspended solids and a biofilter for removing ammonia and organic matter (Chen et al., 1997; Menasveta et al., 2001; Ridha & Cruz, 2001; Skjølstrup et al., 1998; Tal et al., 2006; van Rijn, 1996). RAS with denitrification units (B and C) offer better performance in the removal of nitrite, and nitrate as well as organic matter, compared with the conventional system (A). Particularly RAS C attracts more and more attention (Meriac et al., 2014).
Despite of the fact that most of RAS are equipped with nitrification units, the production of sludges coming from drum-filters, e.g. backwash water, is still significant. Usually these sludges consist of faeces and uneaten fish feed. About 10-30% of the quantity of feed fed to fish is converted to suspended solids (Chen et al., 1997; Timmons & Ebeling, 2007). Research on proper disposal methods for the considerable amount of sludge is considered imperative for preventing any negative environmental impact and maintaining sustainability of RAS (Barak et al., 2003; Chen et al., 1997; Mirzoyan et al., 2010; Timmons & Ebeling, 2007; van Rijn, 1996).
Anaerobic digestion (AD) involves the breakdown of organic material and it has been employed to treat various organic wastewaters, municipal sludges and industrial organic wastes, as well as agricultural residues because of the potential of stabilization of waste streams, bioenergy recovery (biogas production), and sludge reduction (Gujer & Zehnder, 1983).
Nowadays, with the increase in crude oil prices, the potential looming of the exhaustion of fossil fuel reserves, as well as the increased attention on the green-house gases emissions, much more sustainable, cost-effective and environmentally friendly
energy sources are needed to be developed and applied (Wee Shen Lee et al., 2013). Bioenergy recovery from organic wastes, such as the production of biogas that mainly consists of CH4and CO2, is considered an important alterative approach, contributing to our future energy demand. Therefore, there is a growing interest to recover bioenergy from organic-rich waste streams. Produced biogas from wastes can be used to generate electricity and also as heat source for warming, or cooking in rural areas (Madsen et al., 2011).
However, only few studies are available on AD of sludge from marine/brackish RAS, as shown in Chapter 4. Little information on the application of AD in full-scale RAS, especially marine and brackish RAS, is reported in Chapter 2. Moreover, the limited literature data shows unstable performance of batch anaerobic digesters treating the salty sludges, in terms of low COD to methane conversion ratio and unstable COD removal efficiency (Gebauer, 2004; Zhang et al., 2013). The possible causes are discussed in Chapter 2, and include an improper inoculum employed in batch studies and high salinity levels.
1.2 Objectives, scope of the current research and problem definition
Based on information obtained from literature reports on AD of sludges from marine/brackish RAS, the main focus of this thesis is oriented to enhanced concentration of RAS sludge using coagulation, followed by an improved biomethane yield and phosphorus release from the salty sludge during AD providing the possibilities to not only achieve sludge reduction, but also to recover bioenergy and phosphorus.Therefore, the following 5-fold research objectives are formulated:
Firstly, to examine the performance of organic/inorganic coagulants under brackish conditions and to better understand the possibility of applying coagulants in marine/brackish RAS, which is addressed in Chapter 3. In this Chapter, potential impacts of inorganic coagulants on the digestibility of the sludges produced when they are employed to concentrate the waste streams from the marine/brackish RAS are also investigated;
Secondly, to assess specific methanogenic activity (SMA) and phosphatase activity (PA) that is related to phosphorus release from particulate organic matter as well as the occurrence of extracellular polymeric substance (EPS) and soluble microbial products (SMP) of anaerobic sludge fed with salty sludge under high ionic strength condition;
Thirdly, to examine whether addition of compatible solutes, such as trehalose and glycine betaine, could potentially enhance methane production rates and phosphorus release during AD of the salty solids waste;
Fourthly, to investigate the possibility of employing a long term adapted high salinity inoculum to enhance the digester performance in terms of specific methane yield under elevated organic loading rate (OLR), which is addressed in Chapter 6. In this Chapter, the influence of operational parameters, i.e. OLR and sludge retention time (SRT) on the changes of bio-community was observed;
Figure 1.1 Scheme of RAS with three configurations, A-Conventional RAS, B and C-RAS with different positioning of denitrification (sludge disposal unit shall be replaced by digester)
Fifthly, to examine struvite precipitation under the brackish conditions, such as estimation of the thermodynamic solubility product of struvite under the high ionic strength condition, which is limited in literature.
The first research question is whether organic and/or inorganic coagulants can be used to treat backwash water from marine/brackish RAS. This is examined in Chapter 3. The driving force of this question is the lack of information in literature on the performance of organic/inorganic coagulants in treating wastewater from marine/brackish RAS. Moreover, the effluent from the coagulation/flocculation units offers the possibility to be reused in the RAS. The concentrated streams could be further treated using anaerobic digestion. Therefore, the RAS system might be more compact and environmentally friendly via integrating separation/concentration methods such as coagulants and/or anaerobic digestion into the RAS since huge sludge storage and/or treatment units, e.g., ponds, are not needed anymore.
The second research question is whether addition of inorganic coagulants affects the biodegradability of the salty sludge (addressed in Chapter 3) and how it affects the specific methanogenic activity and phosphatase activity of anaerobic sludge (as part of Chapter 4).
Inorganic coagulants FeCl3and polymeric aluminum sulfate (PAS) were dosed to treat backwash water from the salty backwash water. Subsequently the second aim of this research is to investigate the potential effects of inorganic coagulants FeCl3and PAS on the biological degradability of the sludge in terms of biochemical oxygen demand (BOD5) and biochemical methane potential (BMP). The results of BMP tests showed that FeCl3presents less negative effects on BMP, compared to PAS. Thus, further investigation was conducted on the effects of FeCl3on the anaerobic process, i.e. the anaerobic sludge, in terms of specific methanogenic activity and phosphatase activity. Generally, high salinity waste streams present major challenges to biological processes. It has been reported that the relatively low specific methane yield of sludges during the anaerobic digestion of salty sludge is linked to the high salinity in the sludge (Gebauer, 2004). Moreover, it has been reported that compatible solutes enhanced the biomethane potential from organic waste with high salt content (Oh et al., 2008). Therefore, addition of compatible solutes, i.e., glycine betaine and trehalose, was employed to examine whether dosing of the compatible solutes could enhance BMP of the salty sludge.
Based on the outcome of the second research question, the third one is derived, i.e. how compatible solutes affect SMA and phosphatase activity of anaerobic biomass treating salty sludge and whether they can enhance the activities. This research question is addressed in Chapter 5.
The forth research question is to examine whether long term high salinity adapted inoculum would accelerate stable performance of the digester treating sludge with high salinity at elevated OLRs. Meanwhile, it is also to observe the effects of the varying OLR on the distributions of microorganisms at phylum and genus levels. The answer of this question is elaborated in Chapter 6.
Long-term adapted saline inoculum might be conducive for improving the performance of anaerobic digesters treating salty sludge from marine/brackish RAS. Therefore, an inoculum harvested from an anaerobic digester of a marine fish processing factory located in Spakenburg, the Netherlands, was used to inoculate our digester shown as in Figure 1.2. Moreover, to date, the phylum and/or genus distributions of bio-community of anaerobic reactors treating sludge from marine/brackish RAS has not been reported in literature. Understanding the change of the bio-community of the anaerobic reactors at phylum and genus levels could be helpful in better explaining performance of the anaerobic digesters. Therefore, the investigation of the biodiversity of the anaerobic reactor treating the salty sludge was carried out.
The fifth research question aims to investigate struvite precipitation under brackish conditions in terms of thermodynamic solubility product and particle size distribution of struvite. The question is elaborated in Chapter 7.
It is reported in literature that concentrated sludge from marine/brackish RAS contains a relatively high content of phosphorus, compared to municipal sludges (Chen et al., 1997). During AD, ammonium and phosphorus can be released. Therefore, in the digesters, magnesium ammonium phosphate (struvite) precipitation can occur due to the high concentrations of released ammonia and phosphate and the natural presence of magnesium under brackish conditions. However, ionic strength influences the activity coefficients of ions, in particular those of ions with high charge such as PO43-and Mg2+. Therefore, it is of great interest to investigate struvite precipitation under brackish conditions. Thus in Chapter 7, the thermodynamic parameters of struvite formation under the brackish condition, such as thermodynamic solubility product and the enthalpy of struvite formation reaction as well particle size distribution of struvite were estimated and analyzed, respectively.
Figure 1.2 The anaerobic digester used in this study
1.3 Outline of the thesis
This thesis presents studies on: 1) separation/concentration of saline backwash waters from a brackish RAS using coagulants, 2) methods to enhance anaerobic digestion, 3) methane production of salty waste from a brackish RAS and 4) the change of bio-community in the digester as well as 5) struvite precipitation under the brackish condition. The outline of the thesis is summarized in Figure 1.3.
Chapter 2 reviews the literature on the anaerobic digestion of sludges from
marine/brackish RAS and discusses potential causes responsible for the low methane yields from the salty sludges. Moreover, some approaches to the corresponding causes. In Chapter 3, an investigation on the performance of a few organic coagulants and 2 inorganic coagulants including FeCl3 and PAS in treating backwash waters from a brackish RAS was conducted. The tested organic coagulants/flocculants were supplied by Nalco, including organic coagulants (product number (PN): 8103+ and 8108+), a cationic flocculant with very high charge and low molecular weight (PN: 71429), cationic flocculants with high charge and high molecular weight (PN: 71437, 9909 and 9916), a nonionic flocculant with high molecular weight (PN: 71771) and an anionic flocculant
with high charge and very high molecular weight (PN: 71605). In addition, some biological tests were carried out to evaluate the potential effects of addition of inorganic coagulants on biological digestibility of sludges from the brackish RAS located in Zeeland, the Netherlands. Chapter 4 evaluates the effects of salinity and an inorganic coagulant (FeCl3) on the SMA, EPS, SMP and also phosphatase activity of anaerobic sludge fed with the sludge from the brackish RAS. Chapter 5 examines whether addition of compatible solutes, i.e., glycine betaine, trehalose and potassium chloride could enhance SMA, phosphatase activity and phosphorus release of the brackish RAS sludge during anaerobic digestion. In Chapter 6, the performance of a digester fed with sludge from a brackish RAS was investigated during 400 days of operation. In addition, biodiversity of micro-community in the digester was examined. Chapter 7 addresses struvite precipitation under the brackish condition, mainly including estimation of thermodynamic parameters of struvite, crystal size distribution, and composition of struvite. Chapter 8 presents a general discussion of the results and the implications of this current research work.
1.4 References
Aboutboul, Y., Arviv, R., Van Rijn, J. 1995. Anaerobic Treatment of Intensive Fish Culture Effluents - Volatile Fatty-Acid Mediated Denitrification. Aquaculture, 133(1), 21-32.
Barak, Y., Cytryn, E., Gelfand, I., Krom, M., van Rijn, J. 2003. Phosphorus removal in a marine prototype, recirculating aquaculture system. Aquaculture, 220(1-4), 313-326.
Chen, S.L., Coffin, D.E., Malone, R.F. 1997. Sludge production and management for recirculating aquacultural systems. Journal of the World Aquaculture Society, 28(4), 303-315.
Gebauer, R. 2004. Mesophilic anaerobic treatment of sludge from saline fish farm effluents with biogas production. Bioresource Technology, 93(2), 155-167.
Gujer, W., Zehnder, A.J.B. 1983. Conversion Processes in Anaerobic-Digestion. Water
Science and Technology, 15(8-9), 127-167.
Madsen, M., Holm-Nielsen, J.B., Esbensen, K.H. 2011. Monitoring of anaerobic digestion processes: A review perspective. Renewable & Sustainable Energy
Reviews, 15(6), 3141-3155.
Menasveta, P., Panritdam, T., Sihanonth, P., Powtongsook, S., Chuntapa, B., Lee, P. 2001. Design and function of a closed, recirculating seawater system with denitrification for the culture of black tiger shrimp broodstock. Aquacultural
Engineering, 25(1), 35-49.
Meriac, A., Eding, E.H., Schrama, J., Kamstra, A., Verreth, J.A.J. 2014. Dietary carbohydrate composition can change waste production and biofilter load in recirculating aquaculture systems. Aquaculture, 420, 254-261.
Mirzoyan, N., Tal, Y., Gross, A. 2010. Anaerobic digestion of sludge from intensive recirculating aquaculture systems: Review. Aquaculture, 306(1-4), 1-6.
Oh, G., Zhang, L., Jahng, D. 2008. Osmoprotectants enhance methane production from the anaerobic digestion of food wastes containing a high content of salt. Journal
of Chemical Technology and Biotechnology, 83(9), 1204-1210.
Ridha, M.T., Cruz, E.M. 2001. Effect of biofilter media on water quality and biological performance of the Nile tilapia Oreochromis niloticus L. reared in a simple recirculating system. Aquacultural Engineering, 24(2), 157-166.
Skjølstrup, J., Nielsen, P.H., Frier, J.O., McLean, E. 1998. Performance characteristics of fluidised bed biofilters in a novel laboratory-scale recirculation system for rainbow trout: nitrification rates, oxygen consumption and sludge collection.
Aquacultural Engineering, 18(4), 265-276.
Tal, Y., Watts, J.E.M., Schreier, H.J. 2006. Anaerobic ammonium-oxidizing (anammox) bacteria and associated activity in fixed-film biofilters of a marine recirculating aquaculture system. Applied and Environmental Microbiology, 72(4), 2896-2904. Timmons, M.B., Ebeling, J.M. 2007. Recirculating Aquaculture. Northeastern Regional
Aquaculture Center, Ithaca, New York.
van Rijn, J. 1996. The potential for integrated biological treatment systems in recirculating fish culture - A review. Aquaculture, 139(3-4), 181-201.
Wee Shen Lee, Adeline Seak May Chua, Hak Koon Yeoh, Ngoh, G.C. 2013. A review of the production and applications of waste-derived volatile fatty acids.
Zhang, X., Spanjers, H., van Lier, J.B. 2013. Potentials and limitations of biomethane and phosphorus recovery from sludges of brackish/marine aquaculture recirculation systems: A review. J Environ Manage, 131, 44-54.
phosphorus recovery from sludges of brackish/marine
aquaculture recirculation systems: A review
This chapter is based on
Zhang, X., Spanjers, H. and van Lier, J.B. (2013) Potentials and limitations of biomethane and phosphorus recovery from sludges of brackish/marine aquaculture recirculation systems: A review. J Environ Manage, 131, 44-54.
Abstract:
Marine/brackish aquaculture (mariculture) is regarded as a promising approach to fulfil the worldwide rapidly increasing fish demand. However, environmental impact of marine aquaculture is an increasing constraint and regulations on discharge of aquaculture systems are becoming more stringent. The current constraints, however, could be addressed by adopting marine recirculation aquaculture systems (RAS). RAS are highly efficient in production of fish and seafood and cause comparatively less negative environmental impacts because of their low- or zero-emission of wastewater, compared with conventional pond systems and marine based cage systems.
Although the amount of waste from solids-removal units of brackish/marine RAS is relatively small, the produced waste is still perceived as a constraint for sustainable development of brackish/marine RAS. Appropriate disposal of sludge or waste from brackish/marine RAS is of great importance for widespread acceptance and implementation.
Anaerobic stabilization of RAS wastes is considered a potential cost-effective methodology to achieve effective sludge reduction and biogas production. Therefore, an overview is given of studies on anaerobic digestion of sludge from brackish/marine RAS. Several researchers have shown that specific methane production (SMP) of anaerobic digestion of the sludges from brackish/marine RAS is relatively low, 0.020-0.184 CH4 m3 (STP)/kg COD removed. The possible reasons for the low methane production can be attributed to applied experimental set-ups, improper inoculum and short test durations, and are partly related to the characteristics of sludges, including higher fractions of lipids and proteins, low TS contents, high salinity and accumulated long chain fatty acids (LCFA). This review also evaluates the potentials and limitations for phosphorus recovery from the waste streams. Additionally, corresponding approaches to enhance specific methanogenic activities are proposed.
2.1 Marine/brackish aquaculture recirculation system and waste
valorisation
A striking decrease in marine biodiversity has been connected to intensive fishing, environmental pollution and marine and coastal habitat destruction, possibly exacerbated by climate changes. Meanwhile, the consumption of fish has doubled since the early 1970 and is bound to continue to mount with the growth of population, increase in income and rapid urbanization in developing countries (Naylor & Burke, 2005). In order to meet the growing fish demands, aquaculture is being recognized as the most promising and feasible means to provide sufficient fish and seafood (DeVoe, 2000; Naylor & Burke, 2005).
Notwithstanding the significant potential of aquaculture to relieve the pressure from the seafood demand, traditional methods, like earthen ponds, raceway aquaculture and net-pen aquaculture, have their ecologically adverse effects. The three main negative impacts of conventional aquaculture widely discussed in the literature include the salty wastewater emissions from fish farms that contaminate local environments, the escapees which might cause a big loss of gene pool of fish via interbreeding, and the epidemics and parasites which could spread between farmed fish and wild stocks (Gelfand et al., 2003; Naylor & Burke, 2005; Naylor et al., 2005). In addition, traditional marine aquaculture demands a great exchange ratio of fresh seawater, coupled to large footprint requirements (Ackefors & Enell, 1990; DeVoe, 2000; Gelfand et al., 2003; Naylor & Burke, 2005). In contrast, brackish/marine recirculation aquaculture system (RAS) use 90-99% less fresh water, and less than 1% of land, compared with the traditional aquaculture systems. Typical RAS consist of culture units such as fish tanks, solids separation units such as drum-filter commonly used, and also nitrification units-i.e., ammonia removing biofilters (Chen et al., 1997; Timmons & Ebeling, 2007; van Rijn, 1996). The effluent from the culture units goes to solid separation units, then the main flow is directed to the nitrifying biofilters, and after the biofilters and optionally after sterilization and aeration the effluent flows back to culture units. This is the major loop of RAS. Meanwhile, another relatively concentrated stream from solid separation units is produced. Therefore, it is apparent that RAS offer controlled wastewater treatment and waste management, which make RAS much more compatible with local environments. Furthermore, the advantages of RAS include stable product growth rate in a controlled condition and a predictable harvesting schedule as well (Timmons & Ebeling, 2007). Hence, with the demand of sustainable development of mariculture and the more stringent regulations on emissions, marine RAS, with salinity levels exceeding 30 g/L, and brackish RAS, with salinity levels ranging from 0.5 g/L to 30.0 g/L, seem to be the
most optimal approach (Barak et al., 2003; DeVoe, 2000; Timmons & Ebeling, 2007). (Note that salinity values from literature reported in this thesis are assumed to be based on the standard conductivity measurements (APHA, 2005) and here expressed in g/L).
However, the relatively small production of concentrated waste mainly containing organic solids from RAS needs to be treated. Most authors (Mirzoyan et al., 2010; Timmons & Ebeling, 2007) report that solid waste or sludge from RAS is comprised of partially stabilized fish faeces, and a small percentage of uneaten feed. Production and characteristics of sludges from RAS vary widely, but mainly depend on the type of culture system, feed rate, feed type, the size and type of species cultivated and the water treatment process, because the excretions might be partially degraded before they reach the storage units of sludge, such as off-line settling basins (Chen et al., 1997; Mirzoyan et al., 2010). Furthermore, aquaculture sludge contains 1.4-2.6% of total solids (Timmons & Ebeling, 2007) with percentage of volatile organic matter ranging from 50 to 92% (Mirzoyan et al., 2010), and a higher nitrogen and phosphorus content than domestic sludge. Total phosphorus is typically 1.3% of dry solid mass, in contrast with 0.7% of total phosphorus in typical domestic sludge (Chen et al., 1997; Timmons & Ebeling, 2007). Moreover, the major form of biochemical oxygen demand (BOD), chemical oxygen demand (COD) and total Kjeldahl nitrogen (TKN) is particulate. On the contrary, for phosphorus species in sludges from RAS, Chen et al. (1997) reported that most of phosphorus exists in the dissolved form, i.e. orthophosphate, rather than in particulates. Last but not least, sludges from marine or brackish RAS have another specific characteristic, i.e., a high salinity ranging from 2.5- 35 g/L, as shown in Table 2.1, which causes difficulties in the biological treatment of such sludges.
For the treatment of concentrated sludge, clarification commonly is used to concentrate sludge further, and after settling, sludge could be applied as fertilizer for field crops, or be processed further before land application (Chen et al., 1997; Timmons & Ebeling, 2007). Further treatment approaches to stabilize sludges include anaerobic lagoon, aerobic lagoon, aerobic digester, anaerobic digesters and composting. The aerobic systems require aeration facilities which raise the costs. Additionally, for land application of sludge directly or after disposal, transfer costs to fields and odor emissions are always limiting the direct application of the sludge as a fertilizer (Chen et al., 1997). Particularly sludges from brackish and marine RAS containing higher concentrations of salts cannot be applied to fields because of the potential salinization and/or sodification of soil, groundwater and local surface water bodies caused by infiltration (Chen et al., 1997; Mirzoyan et al., 2010; Timmons & Ebeling, 2007). Furthermore, the salty sludge ought not to be discharged into the sewerage system because this may disturb the
normal operation of wastewater treatment plants due to the high contents of salts. Thus far, in most cases, the sludges from brackish or marine fish farms are disposed either by land application or by discharge into sewerage systems (Chen et al., 1997; Sharrer et al., 2007; van Rijn, 1996).
Anaerobic digestion (AD) has been widely applied to dispose of highly concentrated organic wastewaters, municipal sludges and industrial organic wastes, as well as agricultural residues because of the potential of biogas generation and sludge reduction. However, little information of the application of AD in full-scale RAS, especially marine and brackish RAS, has been reported (Mirzoyan et al., 2010). Tal et al. (2009) estimated that 1000kg fish production from the RAS generates about 375 kg of dry organic solid waste, based on a food conversion ratio of 1.5 and 75% feed uptake by the fish. Only 10.5 kg or 2.8% of the original dry organic waste will maintain after AD of the organic particulate sludge from the marine RAS (Tal et al., 2009). Furthermore, AD of sludge from marine RAS can also produce volatile fatty acids (VFA) that can be used as carbon source in the RAS denitrification units. Moreover, phosphorus release from particulate sludge occurs in AD, which provides reactive phosphorus that can be recovered from the digestate as phosphate salts, such as struvite (MgNH4ÂPO4Â6H2O). Apparently, AD
of sludges from RAS seems to be a potential means to achieve reduced emission from RAS, eventually leading to almost “zero-discharge RAS”. However, AD processes, particularly with anaerobes non-adapted to saline condition, can be inhibited by high salinity. Therefore, high salinity may be the crucial factor that limits the application of AD in treatment of waste streams from brackish/marine RAS. A review of studies on AD in the disposal of sludges from brackish/marine RAS is needed to further examine limitations on bioenergy recovery from brackish/marine RAS and the potential application of AD in those systems.
This review summarizes literature reporting research on anaerobic/anoxic biological methods to dispose of concentrated streams, such as backwash waters and sludges from marine and brackish RAS. Particular attention is paid to possible causes leading to lower methanogenic activities in AD of sludges from marine and brackish RAS. The possible causes are examined and listed in the sequence: inappropriate inocula, sludge composition (e.g., macromolecules, TS and sulfate), high salinity levels of sludges, particularly high sodium concentrations, ammonia generated in the digestion of nitrogenous compounds, accumulation of long chain fatty acids (LCFA) and synergistic effects of magnesium ions and calcium ions on methanogenic activity. We also evaluated the phosphorus accumulation problem in RAS, resulting from lack of
phosphorus removal units. Afterwards, feasible approaches to improve methanogenic activities in AD of sludges from marine and brackish RAS are proposed.
2.2 Advantages of AD of wastes from marine/brackish RAS
AD is a reductive conversion and/or biodegradation of organic matter and/or inorganic matter (e.g. sulfate) by facultative and obligate anaerobes in the absence of molecular oxygen (O2) and nitrate, resulting in the generation of CH4, CO2 and/or H2S (Speece, 2008; Tchobanoglous et al., 2003). The AD of complex wastes, such as those involving particulate or suspended matter, comprises a series of sequential biochemical processes, as shown in Figure 2.1: disintegration, hydrolysis, fermentation, acetogenesis and methanogenesis (Gujer & Zehnder, 1983; Khanal et al., 2008; McCarty & Smith, 1986).
AD is a relatively slow degradation process, mainly resulting from a slow growth rate of methanogens and low hydrolysis rate with particulate organic matter (Gujer & Zehnder, 1983; van Lier, 2008). Despite the drawback, AD offers the following striking advantages, compared to direct land application and composting (Mata-Alvarez et al., 2000; van Lier, 2008), which are most commonly used in disposal of sludges from fish farms (Chen et al., 1997).
1. Significant reduction in space requirement;
2. Reduction of excess sludge and meanwhile bioenergy recovery; 3. High applicable organic loading rates;
4. More compatible with wastewater treatment units of RAS, compared with other conventional methods for disposal of concentrated wastes;
5. Lower risk of further contamination caused by sludges with high contents of salts, such as infiltration to soil and groundwater and odor from sludges, compared to direct land application;
6. Small scale AD applications allow decentralization of sludge treatment for fish farms;
7. Anaerobes could survive under non-feed conditions for long periods, without serious exacerbation of the activity.
Figure 2.1 Anaerobic Digestion Process of Composite Particulate Material (After Gujer and Zehnder, 1983)
2.3 AD of sludges from brackish/marine RAS
2.3.1 Anaerobic and aerobic reactors in disposal of backwash waters and sludge from brackish and marine RAS
In sludge treatment of marine/brackish RAS, 5 types of biological reactors (Figure 2.2) have been reported, including anaerobic completely stirred tank reactors (CSTR) (Figure 2.2A) (Gebauer, 2004; Gebauer & Eikebrokk, 2006), upflow anaerobic sludge bed reactors (UASB) ( Figure2.2B) (Mirzoyan & Gross, 2013; Mirzoyan et al., 2008; Mirzoyan et al., 2010; Tal et al., 2009), and anoxic tank plus aerobic membrane bioreactor (MBR) (Figure 2.2C) (Sharrer et al., 2007), waste stabilization ponds (WSP) (Mirzoyan et al., 2012), and anaerobic sequencing batch reactor (ASBR) (Figure 2.2A) (Luo et al., 2013a).
Figure 2.2 Biological reactors used for sludge digestion in RAS: A-Completely Stirred Tank Reactor (CSTR) and Anaerobic Sequencing Batch Reactor (ASBR); B-Upflow Anaerobic Sludge Blanket Reactor (UASB); C-Membrane Bioreactor (MBR)
Table 2.1 summarizes the characteristics of biological reactors in disposal of sludges from brackish/marine RAS, including characteristics of sludges, reactor types, removal efficiencies of COD, sludge reduction percentages, methane percentages in biogas, as well as methane yield. Table 2.2 summarizes some operational parameters of marine and brackish RAS and the bioreactors used in these systems, including salinity, cultured fish species, feed rate and volume of digesters, hydraulic retention time (HRT), organic loading rate (OLR), temperature, duration of experiment and methane production.
Table 2.1 demonstrates that UASB and MBR showed high performances in terms of COD removal in treating sludges from marine/brackish RAS with low TS, whereas WSP showed poor performance in removal of organic matter, compared to UASB (Mirzoyan et al., 2012). Moreover, the costs of MBR systems are generally higher, compared to UASB and CSTR, which might limit its application in sludge/wastewater treatment from brackish/marine RAS. Hence, UASB is a feasible approach to achieve carbon valorization for sludge with low TSS. In addition, for disposing sludges with high TS, particularly high TSS, which is not suitable as feed for UASB and MBR, CSTR is the
appropriate approach due to its relatively high COD removal efficiency in treating concentrated waste stream.
Literature on AD of sludges from brackish/marine RAS is still limited and no application of AD in full-scale brackish and marine RAS has been reported. Additionally, biogas production is low in most of the studies on AD of sludges from brackish and marine RAS. Hence, the causes limiting biogas recovery need to be further elucidated.
T
a
ble 2-1 Chara
cteristics
and biogas production of
sludg es w ith differ ent sali nity f rom bra cki sh/ma rine R A S T y pe of W a s te o r W a s tew ater s T y pe of reac tor Sal in it y (g/L) TS (%) VS (w t% of T S ) TN (m g/L) TP (m g/L) N:P mo la r rati o Diges ti on ef fi c ienc y (%COD) Slu dg e reduc tio n (%) Meth ane (% bio gas ) Yie ld of 0HWK DQH /ÂJ -1 COD added )) W a s te f ro m m a ri ne RAS (T al et a l., 2009) UAS B 15-17 3 (w/v ) -1.44 - 80 62 0.45 a S lud ge fro m sa lin e fi s h f a rm ( G ebauer , 2004) CST R 35 8.2– 10.2 49.8 – 54.1 2442- 3043 1350 – 1683 4.01 36.7-5 5.2 - 48.9-5 4.1 0.11 4-0.1 84 8.0-13. 7 a ˅ 17.5 4.1- 5.1 - - - - 6 0 - 5 7 .6 0.15 4 ˄ 5. 2 a ˅ Sal ty s ludg e( G eba uer & Eik ebr ok k , 2006) CST R 1 4 6.3– 12.3 78.6 – 86.9 5456- 1064 1 1424 – 2780 8.48 44-54 - 59-61 0.14-0. 15 (24.8-25 .7 a ) S lud ge fr o m br ac k is h aquac ul tur e (Mir z o y a n et a l., 2008) UAS B 2 -5 1.49- 2.16 (w /v ) 56.1- 75.6 - 3-10.5 60 b 35-70 30-60 <0. 012 B a ck w a sh fl ow fro m fr es h RAS ( S har re r et a l., 2 007) AT + AM BR c 0 0. 17 e 81.7 e - 68.4 57.2 99.9 9 2.65 - - AT + AM BR c 8 i 0. 17 e 83.9 e -67.3 38.7 99.9 6 3.85 -AT + AM BR c 32 i 0.08 e 85.1 e - 50.7 19.2 99.9 6 5.85 - - S lud ge fr o m br ac k is h aquac ul tur e (Mir z o y a n et a l., 2010) UAS B - 0.4 - - - - 92-98 - 4-53 0.00 004-0.00 36 1 .4 - - Com par able t o GE d S lud ge fr o m br ac k is h f a rm (Mir z o y a n et a l., 2012) W S P 2 .5 0.12 e 7 3 - - - - 2 2 f - - S lud ge fr o m br ac k is h f a rm (M ir z o y a n & G ros s , 2013) UAS B 2.6 ±0 .1 0.38 ±0 .0 2 -- -- - 97.7-9 9.6 - 15, 4 , 53 , 2 0.01 5, 0. 005 , 0.07 5, 0. 001 h
S lud ge fro m a R A S (Luo et al ., 2 013 a) AS BR 0-10.5 i - - 2010 0-2430 0 g -- - 38-73 88-98 28-57 0.01 3-0.0 22 D 8Q LWRI WK HP HWKDQH SU RG XF WLRQ U DWH LV /Â/ -1 s lu dge b. T O C r e m oval ef fi c ienc y c . A T + AM B R : An ox ic T ank + Aer at ed MBR d. G E : CH 4 y ie ld s im ilar to t he m ethane pr oduc ti o n ( G e bauer & Eik ebr ok k , 2006) e. T S S an d V S S i n th is c a s e f. or g anic m atter r e m oval ef fi c ienc y g. T otal k jeldahl ni tr oge n h . C a lc u lated on th e bas is of biog as pr od uc tio n an d m ethane c o nc entr at ion i n bio gas ( M ir zo y a n & G ros s , 2013) i. the s a lin it y of the s lud ges was ac hi ev ed b y add it ion of NaCl ( S har re r et al ., 2 0 0 7 ) and s e a s a lts ( Luo et a l., 201 3a) . T a ble 2-2 A q u aculture Re circul ation S ystems w ith A D of sludges a re c a lc ul ate d v a lues bas e d on CO D: SO 4 2- , SO 4 2-c o n c entr ati on, wor k ing v o lu m e of UAS B ( 6L) , and f e e d ing ra te to UA SB ( 4 00 m L /da y ) f ro m the s tud y ( Mir zo y a n et a l., 200 8) , in whic h CO D is es tim ated f rom T O C RAS Ty p e Fi sh T y pe Slu dg e d iges ti on ty pe T (°C) Vol um e of diges ter (m 3 ) 2/5 J&2'Â / -1 Âda y -1 ) HRT (day s) Auth or s Mar in e Salm on CST R 3 5 0.01 5 1.24-3. 12/ 1.1 - G ebau er ( 200 4) Br ac k is h Salm on s m olt C ST R 3 5 0 .01 5 2.9/3 .1 55/2 4 0 G ebau er an d E ik ebr ok k (2006) Br ac k is h T ilapi a U AS B 2 5 0 .00 6 23-60 a 1 5 M ir z oy an e t al ( 2 00 8 ) Br ac k is h Str ip ed bas s U AS B 11- 30 -Mir zo y a n et a l.( 20 10) F re s h Rain bo w T rout Aer at ed M B R - - - 40.8 Shar re r e t a l.( 20 07) Br ac k is h Sea bas s U AS B 2 6 0.5 - 10.4 T al et al . ( 2 0 09) Br ac k is h Str ip ed bas s W SP 11- 33 - 50-60 Mir zo y a n et a l.( 20 12) Br ac k is h T ilapi a U AS B 11- 27 0.00 8/0. 006 - 8/6 M ir z oy an a n d Gro ss (2013) Fre sh Sc or tum bar c o o AS BR - 0.00 4 0.12-0. 41 20 Luo et a l. ( 2 0 1 3 a ) 20
2.3.2 Inocula as a key factor for saline wastewater and sludge treatment
Suitable inocula and acclimation strategies are imperative for a successful start-up and stable operation of anaerobic reactors dealing with wastewaters particularly containing high salt contents. Using anaerobes from saline environments for seeding the system is an alternative to shorten the fresh-water inoculum adaption period. A number of researchers (Aloui et al., 2009; Ismail et al., 2008; Sharrer et al., 2007) have been using acclimation (Kobayashi et al., 2009; Le Marrec et al., 2007; Oh et al., 2008; Rontein et al., 2002) to adapt the bacteria to saline environments and have obtained some good results. Moreover, application of halophilic microorganisms in biological treatment of saline wastewater has also been studied experimentally by several researchers (Choi & Park, 1999; Hinteregger & Streichsbier, 1997; Jeison et al., 2008; Kargi & Dincer, 2000; Woolard & Irvine, 1995), and the approach of seeding anaerobic reactors with halophilic sludge treating saline wastewaters seems to be advantageous (Abou-Elela et al., 2010; Kargi, 2002; Kargi & Dincer, 2000). Aspe et al. (1997) demonstrated that a marine inoculum adapted much better and faster at 37 ºC in digesters for fishery wastewater with salinity of 25.6-31.5 g/L than fresh pig manure as an inoculum.
Table 2-3 Inoculum adopted in the studies on the disposal of sludge from brackish/marine RAS
Inoculum source Biological reactor type
Experimental
period (day) Reference An anaerobic digester originally seeded with a
mixture of digested municipal sludge and cow manure
CSTR 440 Gebauer (2004)
An anaerobic digester fed with saline fish farming sludge with salinity of 35 g/L and operated for 4 months (Gebauer, 2004)
CSTR 155 Gebauer and Eikebrokk (2006) 4 months incubation of brackish aquaculture
sludge without addition of inoculum UASB 120
Mirzoyan et al. (2008)a
- UASB 131 Tal et al.(2009)
Sludge from a UASB running for 120 days UASB 335 Mirzoyan and Gross (2013) Digestate of a municipal sewage plant ASBR 165 Luo et al.
(2013a)
-Not available
a
This information was obtained via personal communication with Mirzoyan in March 2014.
In the studies of AD of sludges from marine/brackish RAS, the choice of inocula obviously plays an important role. Specific methane yield (SMY) in the reported studies of Gebauer (2004), Gebauer and Eikebrokk (2006), Mirzoyan et al. (2008), Mirzoyan et al. (2010), Mirzoyan and Gross (2013),Tal et al. (2009) and Luo et al. (2013a) is low, 0.114-0.184, 0.140–0.154, <0.012, 0.00004-0.0036, 0.001-/ÂJ-1COD added, 0.45 /Â/-1 sludge added and 0.013- /ÂJ-1 COD added, respectively. In the studies
conducted by Gebauer (2004) and Gebauer and Eikebrokk (2006), high methane yields probably resulted from the long operation duration and the adapted inocula. The operational period of the study (Gebauer, 2004) was 440 days (Table 2.3), during which a higher degree of acclimation and adaptation of bacteria can be expected. Additionally, Gebauer and Eikebrokk (2006) inoculated a CSTR with sludge cultured in saline conditions (35 g/L) for 4 months (Gebauer, 2004) and operated for 155 days. The two studies have shown relatively higher SMY, 0.114-0.184 and 0.140– /ÂJ-1 COD added, than the others. In the study (Mirzoyan et al. 2008), the reactors were started with brackish aquaculture sludge with no inocula additions and a 4 month period was given to achieve acclimatization of methanogenic community and stabilization of operation before the 120 day operational data was collected and reported (additional personal communication). Luo et al. (2013a) used digestate of a municipal sewage plant as the inoculum of ASBRs, which was not cultured under condition with high salinity levels. Furthermore, Rinzema et al. (1988) concluded from an investigation of sodium inhibition on acetoclastic methanogens in granular sludge from a UASB reactor that acclimation of Methanothrix (currently renamed to Methanosaeta (Patel & Sprott, 1990)), the predominant acetoclastic methanogens, could not be obtained at higher sodium concentrations within a period of 12 weeks either. Gebauer (2004) presumed that strong inhibition of the anaerobic process by high salinity in a CSTR fed with undiluted sludge with TS of 8.2% to 10.2%, even after being operated 440 days, might be attributable to the inoculum cultured at lower salinity and lacking of tolerant species. Hence, the type of the inoculum and the used short adaptation times might be partially responsible for the rather low methane productions in AD of sludges with salinity levels, ranging in 15-17 g/L (Tal et al., 2009), and 0-10.6 g/L (Luo et al., 2013a).
2.3.3 Effect of contents of macromolecules of sludges on AD
The characteristics of complex wastes, referring to e.g. composition, amounts of proteins, lipids, carbohydrates, fibers, and polarity, solubility, and tertiary structure of their components, are crucial factors in the successful operation of AD (Appels et al., 2010; Lesteur et al., 2010). The biodegradability of lipids and proteins is relatively low among macromolecules, in contrast to most of the carbohydrates, except for hemicellulose and non-degradable organics, such as lignin, since the hydrolysis constants (kh) of proteins and lipids generally are lower, compared with those of
carbohydrates, making hydrolysis the rate limiting step in anaerobic digestion. Therefore, the contents of protein and lipid in sludges directly determine the appropriate sludge retention time (SRT) for anaerobic reactors, such as CSTR, which can be explained by
the equation SRT=ln(S1/S2)/kh, S1 and S2 representing biodegradable substrate contents in influent and effluent, respectively. Additionally, less soluble proteins are even more slowly degraded than those with high solubility; polar lipids are much more easily decomposed than lipids with lower polarity.
Biogas production is closely associated with the characteristics of the material to be digested such as composition of wastewaters or sludges. Robbins et al. (1989) demonstrated that substrate with high contents of proteins could cause inhibition of biogas production, due to accumulation of ammonia and increasing pH. Moreover, in two studies conducted by Gebauer (2004) and Gebauer and Eikebrokk (2006) on AD of sludges from marine systems in CSTR, similar methane productions were achieved, i.e. 0.136 L CH4J&2'DW2/5RIJ&2'Â/-1Âd ay-1 and 0.140 L CH4/g COD at OLR of J&2'Â/-1Â day-1, respectively. However, SRTs were totally different, 24 days and 60 days, respectively. The required longer SRT of 60 days (Gebauer & Eikebrokk, 2006), might be related to the composition of the sludge with higher contents of proteins (60%) and lipids (31%) shown in Table 2.4 and lower contents of carbohydrates (9%), compared to the former work (Gebauer, 2004).
Table 2-4 COD to sulfate ratio and macromolecule content of sludges from brackish/marine RAS COD:SO42- Protein (%VS) Lipid (%VS) Carbohydrate (%VS) References
17-27 29 15 56 Gebauer (2004)
96-297 60 31 9 Gebauer and Eikebrokk) (2006)
218-580 - - - Mirzoyan et al. (2008)
22-55 - - - Tal et al.(2009)
-: Not available
The composition of sludges to be anaerobically digested, particularly fractions of protein, lipid and carbohydrate, has a significant effect on SRT of anaerobic reactors. Higher SRTs are required in anaerobically stabilizing sludges with higher contents of proteins and/or lipids (Alves et al., 2009; Cirne et al., 2007). The composition of sludges from RAS, however, varies with fish species and units of separation and/or concentration of sludges. Hence, in AD of sludges from brackish/marine RAS, characterization of the sludges is of great significance to determine appropriate SRTs (Alves et al., 2009) and to approach the potential methane production from the sludges.
2.3.4 Effect of sludge TS on AD
Generally, sludges from marine/brackish RAS with higher TS content, particularly with high TSS, generate much higher methane yields (Table 2.1). In the studies on treatment of sludges with high TS content ranging between 4.2-5.1% (Gebauer, 2004), 8.4-10.2% (Gebauer, 2004) and 6.3-12.3% (Gebauer & Eikebrokk, 2006), methane production averaged about 0.154, 0.114-0.184 and 0.154 L CH4/g COD added,
respectively. With the digestion of sludge containing 3% solids (Tal et al., 2009), the methane production was very low, around 0.45 L CH4ÂL-1 sludge, compared to the methane yield from the other two studies, 5.2-13.7 L CH4ÂL-1 sludge (Gebauer, 2004) and 24.8-25.7 L CH4ÂL-1sludge (Gebauer & Eikebrokk, 2006)). In the study conducted by Tal et al. (2009), a 20 L-UASB reactor was partially fed with sludge from an anaerobic sludge digestion tank that produced VFA as carbon sources for a DENAMMOX fixed-bed up-flow biofilter. Likely, the VFA harvesting to feed DENAMMOX caused a lower methane yield in the UASB owing to the lowered biochemical methane potential (BMP) of the remaining sludge from the tank. Mirzoyan et al. (2010) reported that the methane yield of a UASB reactor fed with salty sludge (1.4%TS) from a RAS was comparable to that of the study conducted by Gebauer and Eikebrokk (2006). Additionally, the lower methane yields in the two studies (Mirzoyan et al., 2008; Mirzoyan et al., 2010), i.e. less than 0.012 L CH4/g COD added and 0.00004-0.0036 L CH4/g COD added, respectively, might be attributed to the low TS of the substrates, particularly low TSS, 1.11-1.71% and 0.4%. Lower F/M ratio, that is: low COD to sulfate ratio, results in more energy loss via sulfate reducing pathway, which is explained in the next section. Notably, the methane yield of the sludge with the lowest TS (0.4%) (Mirzoyan et al., 2010) was the minimum value in the studies on the AD of sludges from brackish/marine RAS (Gebauer, 2004; Gebauer & Eikebrokk, 2006; Mirzoyan et al., 2008; Mirzoyan et al., 2010; Tal et al., 2009). Additionally, the losses of dissolved methane in the reactor effluent (Gimenez et al., 2012) also gave substantial contribution to the low SMY from the UASB reactors (Mirzoyan & Gross, 2013; Mirzoyan et al., 2008; Mirzoyan et al., 2010; Tal et al., 2009) or ASBR (Luo et al., 2013a). Moreover, the loss of methane should be reduced if full-scale UASB reactors are applied in RAS also because methane is a powerful greenhouse gas (Howarth et al., 2011; Lashof & Ahuja, 1990). In contrast, the loss of dissolved methane from CSTR fed with sludges with high TSS (Gebauer, 2004; Gebauer & Eikebrokk, 2006) is negligible due to the high methane production rate from the highly concentrated sludges. Therefore, concentration and separation of the waste streams from RAS is needed since it is conducive to improving biogas production from the sludges and downsize the anaerobic sludge treatment units.
2.3.5 Effect of sulfate concentrations of sludges on AD
High sulfate concentration in sludge from marine or brackish RAS may reduce methane production since sulfate can act as alternative electron acceptor, generating
H2S. Owing to the higher affinity for acetate and hydrogen as well as the small
always win the competition for the reducing equivalents in a mixed environment (Chen et al., 2008). Moreover, H2S as the product from sulfate reduction possibly may reach the toxicity thresholds for methanogens (Bhattacharya et al., 1996; Mirzoyan et al., 2008; Tal et al., 2009). On the other hand, the high COD to sulfate ratio exceeding 10, which is apparent in marine/brackish RAS concentrated wastes, is favorable for the methanogenic process (Bhattacharya et al., 1996; Mirzoyan et al., 2008). In the reported studies (Gebauer, 2004; Gebauer & Eikebrokk, 2006; Mirzoyan et al., 2008; Tal et al., 2009) the salty sludges from brackish/marine RAS commonly exhibit COD to SO4
2-ratios in the range from 17 to 580, as shown in Table 2.4. Hence, AD of sludges from marine/brackish RAS is not expected to be limited by high concentrations of sulfate.
2.3.6 Effect of major monovalent alkali metal ions, Na+and K+, on biogas
production in brackish and marine RAS
High salinity of waste streams may affect biochemical processes of non-halophilic bacteria mainly via the following mechanisms: (1) dehydration and cell lysis resulting from hyperionic and hyperosmotic stress; (2) non-competitive inhibition of the activity of intracellular and extracellular enzymes, hence affecting the general bacterial metabolism; (3) an adverse physical effect on the cell membranes of bacteria, which causes malfunction of cells (Debaere et al., 1984; Galinski & Truper, 1994; Mahajan & Tuteja, 2005).
The high ionic strengths in wastewaters and/or wastes in most cases are primarily caused by NaCl (Lefebvre & Moletta, 2006), which also is the case in the sludges from brackish/marine RAS shown in Table 2.5, whereas high sodium and chloride concentrations can cause toxicity in anaerobic reactors. Various researchers showed that on a molar basis sodium was the most common inhibitor on activity of bacteria among the cations (Debaere et al., 1984; Speece, 2008). The impact of Na+and dose response on anaerobic conversion has been researched by various authors (Debaere et al., 1984; Kirby et al., 2006; Uygur & Kargi, 2004). Sodium showed moderate and strong inhibitory impacts in anaerobic treatment with non-adapted sludge at 3.5-5.5 g/L and exceeding 10g/L, (Kugelman & Mccarty, 1965; Lefebvre & Moletta, 2006). High sodium concentration, i.e. exceeding 10g/L, drastically reduces methanogenic activity of non-adapted sludge (Dereli et al., 2012; Jeison et al., 2008).
In addition, K+, which is commonly present in sludges from brackish/marine RAS, may perform as one of the best extractants for metals like Ni, Mo, and Zn that are bound to sites of sludges and then induce toxicity to microorganisms (Chen et al., 2008; Speece, 2008), likely leading to low activity of anaerobic sludges. However, the K+
concentrations of the typical brackish/marine sludges (Gebauer, 2004; Gebauer & Eikebrokk, 2006; Mirzoyan et al., 2008) listed in Table 2.5 are not reported to be inhibitory for the anaerobic processes (Chen et al., 2008). Contrarily, K+lower than 400 mg/L may improve the performance of reactors due to its antagonism to Na+, Ca2+, and Mg2+and its functionality as an osmoprotectant for cells (Speece, 2008; Vyrides et al., 2010) .
The high salinity, mainly caused by NaCl, of the sludges from brackish/marine RAS was probably the major cause of lower methane production in AD of sludges from brackish or marine RAS. As discussed above, high salinity distinctly impacts the metabolic activity of anaerobes, such as propionic bacteria (Feijoo et al., 1995) and aceticlastic methanogens (Waldron et al., 2007), leading to VFA accumulation and a potential pH drop, subsequently leading to complete process deterioration. In the study conducted by Gebauer (2004), the methane yield was 0.114-/ÂJ-1COD added at a high salinity level of 35 g/L. Particularly high sodium concentrations of 10-10.5 g/L, could have inhibited the anaerobic process, being responsible for the variable VFA composition in the digestate. COD removal efficiencies during this experiment dropped from 55.2% to 36.7%. Gebauer (2004) demonstrated that the performance of a CSTR treating sludge at a salinity level of 35 g/L was unstable, resulting in high concentrations of propionate compared to the other VFAs. In contrast, low VFA concentrations in the digested sludge, stable methane composition in biogas and high COD removal efficiency of 60% were obtained during digestion of diluted sludge with a salinity level of 17.5g/L in the CSTR. Moreover, Luo et al. (2013a) reported that with the increase in salinity from 0 to 7 g/L by adding sea salts mainly containing NaCl into the influents of the reactor, soluble COD of the effluent from the ASBR increased from 130 mg O2/L to 394 mg O2/L. The authors showed that when the salinity of the reactor was increased further from 8.7 g /L, biogas production ceased.
T a ble 2-5 M a jor ions in sl udges and ba ck w ash w a te r of mar ine/br acki sh R A S Na (m g/L) K (m g/L) Ca (m g/L) Mg (m g/L) Cl (m g /L) Fe (m g/L) Cu (m g/L) Zn (m g/L) Mn (m g /L) Sy ne rgi st ic e ffe ct Auth or s Na/Mg Na/Ca Ca/Mg 1020 0 476 4640 1759 2360 0 -+ + + G ebau er ( 200 4) 5100 238 2320 880 1180 00 -210 73 3852 189 1325 -- - × × + G ebau er an d E ik ebr ok k , (2 006) 2838 222 6220 1150 -176 14 93 63 + + + Mir zo y a n et a l. ( 2 008) 4621 309 2837 724 -164 12 53 19 + + + 3022 246 2782 694 -173 443 137 55 + + + a . s ali ni ty le ve ls ar e es tim ated o n th e bas is of elec tr ic a l c onduc ti v it y -: Not a v a ila b le/n o d a ta ×. T her e is no s y ner gi s tic ef fe c t +: T he s y n e rg is tic ef fe c t ex is ts
In conclusion, salinity, mainly caused by NaCl, is the major cause for the instability and low biogas production of digesters fed with saline sludges from brackish/marine RAS. However, adoption of inocula acclimatized to saline conditions that represent the ion composition of brackish/marine RAS seems to be a solution to achieve fast start-up and maintain stable performance of reactors treating sludges from those systems.
2.3.7 Effect of ammonia on AD
Ammonia is generated during anaerobic degradation of organic nitrogenous compounds such as proteins and amino acids. In mariculture systems, sludges contain faeces and excess feed (Gebauer, 2004; Gebauer & Eikebrokk, 2006), which could produce high levels of ammonia during the anaerobic process. An ammonia level exceeding 4,000 mg NH3-N/L was reported to exhibit an inhibitory effect during AD of cattle manure (Angelidaki & Ahring, 1994). Free ammonia is the most toxic form of ammonia nitrogen to methanogens, as the free unionized ammonia can freely penetrate the cell membrane (Bocher et al., 2008; Khanal et al., 2008).
In their studies on the digestion of sludges from brackish/marine RAS, Gebauer and Eikebrokk (2006) reported inhibition of biogas production by ammonia, which might have a link with the high content of proteins in the sludge, i.e. 60%, as shown in Table 2.4. High concentrations of total ammonia (4,858-7,463 mg NH3-N/L) generated in the AD of proteins in digesters (Nielsen & Ahring, 2007), which were significantly higher than the aforementioned inhibitory ammonia concentrations of 4,000 mg NH3-N/L (Angelidaki & Ahring, 1994), might have inhibited aceticlastic methanogens, causing VFA accumulation reaching 18-28 g/L in the digesters. The acetic acid concentration in the range from 7,870 to 10,140 mg/L in a CSTR operated at total ammonia concentrations of 4,858-7,463mg NH3-N/L and free ammonia concentrations of 134-230 mgNH3-N/L (Gebauer & Eikebrokk, 2006) was much higher than that in a CSTR operated at total ammonia concentrations of less than 1,694 mg NH3-N/L and free ammonia concentrations less than 20 mg NH3-N/L (Gebauer, 2004). Results on accumulating VFA showed that acetotrophic methanogenesis in the CSTR (Gebauer & Eikebrokk, 2006) was inhibited as well by high levels of ammonia. Additionally, the stable and slightly increased pH, i.e. 7.3-7.5, in the CSTR operated in the study conducted by Gebauer and Eikebrokk (2006), compared with a pH ranging from 6.7 to 7.0 in the CSTR studied by Gebauer (2004), might also be related to the high protein contents in the sludge since 1 mole of alkalinity is produced in the generation of 1mole NH4+, neutralizing the increasing acidity. Obviously, free ammonia concentrations increase with increasing pH.
However, under high ionic strength conditions activities of ions decrease, including the activity of ammonia (Ronteltap et al., 2007b). The formation of free ammonia may be weaker under marine condtion than under freshwater conditions. Therefore, anaerobes may exhibit higher tolerance towards total ammonia nitrogen due to the reduced acitivity of ammonia under marine/brackish condition. Thus, the effect of ammonia on AD of slduges from marine/brackish needs further investigation.
2.3.8 Effect of LCFA on AD of sludge
The higher percentage of lipids in the sludge (31%) from brackish RAS (Gebauer & Eikebrokk, 2006), in contrast with 15% in the sludge from marine RAS (Gebauer, 2004), was possibly the cause of LCFA inhibition during AD of sludge from brackish RAS. LCFA are produced during hydrolysis of triacylglycerol lipids. Cirne et al. (2007) reported that with increasing content of lipids in wastes-i.e., in the range from 31% to 47% on the basis of COD, stronger inhibition was observed in digesters and the accumulation of LCFA was the main obstacle for methane production. The instability of AD (Cuetos et al., 2010) and the inhibitory effect of lipids, and LCFA, generated mainly during the hydrolysis of lipids, are generally attributed to their adsorption to the bacterial membrane, likely hampering the transport of substrates and products across the membrane, and the interference of the a-polar chain with the bacterial membrane (Pereira et al., 2005). LCFA even at millimolar concentrations were able to inhibit methanogenesis (Koster & Cramer, 1987). Hwu and Lettinga (1997) reported that 50% inhibition of oleic acid to acetoclastic methanogens occurred in the range of 0.53-2.27 mM at 40ºC. Gebauer and Eikebrokk (2006) presumed that the inhibition from LCFA that occurred during the research was probably linked to the presence of oleic acid which amounted 9% of the LCFA in the sludge, and may derive from herring fats as the main fish feed. In addition, as reported that calcium and magnesium can abate inhibition from LCFA on methanogenic activity (Koster, 1987), LCFA inhibition, however, was still observed by Gebauer and Eikebrokk (2006) in treating sludge with relatively high concentrations of calcium and magnesium, as shown in Table 2.5. Thus, effects of LCFA and their species in the AD of sludges under marine conditions are still unclear, and further investigations are needed to determine the role of LCFA. However, the accumulation of LCFA in the anaerobic reactors may also be relieved by adopting an appropriate SRT.
2.3.9 Effect of alkaline earth ions mainly Ca and Mg on methanogenic activity
Synergistic and antagonistic effects of alkaline earth ions on anaerobic processes in the AD of sludges from marine/brackish RAS probably occur as a result of the abundant
presence of, especially sodium, calcium and magnesium (Table 2.5) (Kugelman & Mccarty, 1965; McCarty & McKinney, 1961). Kugelman and McCarty (1965) conducted an investigation on synergistic and antagonistic effects of sodium, potassium, ammonia, calcium and magnesium in digesters. They reported that at certain concentrations of magnesium, i.e. 1,800 mg Mg/L, sodium and magnesium showed antagonistic effects on AD with increasing sodium concentrations. However, after a certain concentration of sodium, i.e. 2,300 mg Na/L, which corresponded to the magnesium concentration of 1,800 mg Mg/L, the two cations presented synergistic effects on AD processes with increasing sodium concentrations. Furthermore, the synergism was even stronger with the continuing increase in sodium concentrations. Thus, according to that study, synergistic effects of the cations, sodium, calcium and magnesium (Table 2.5), which might affect specific methanogenic activity, can be expected during the digestion of the sludges from marine/brackish RAS (Gebauer, 2004; Gebauer & Eikebrokk, 2006; Mirzoyan et al., 2008; Mirzoyan et al., 2010; Tal et al., 2009).
2.3.10 High phosphorus contents in sludges from RAS
According to the mass balance of phosphorus in a RAS culturing eels (Suzuki et al., 2003), 16% and 19% of phosphorus of the feed ended up in fish biomass and rearing water, respectively. However, the major phosphorus fraction of the feed, i.e. 65%, had been accumulated in sludges. The concentrated biosolid streams (sludge) from backwash flows of RAS contain significantly higher concentrations of phosphorus, averagely 1.3% of the dry solid mass, compared to the typical domestic sludge that contain about 0.7% of total phosphorus (Chen et al., 1997; Mirzoyan et al., 2008; Timmons & Ebeling, 2007). Tables 2.6 and 2.7 show that in backwash waters from drum filters of RAS and concentrated sludge, organically bound phosphorus or particulate phosphorus is the dominant phosphorus form. Phosphorus concentrations expressed as % of TS, in sludges from RAS varies hugely, ranging from 0.2% to 12% (Table 2.7), which is closely related to RAS operational conditions, such as fish species and solids separation methods. With regard to the phosphorus fraction in faeces, uneaten feed and biomass of sludges from RAS, however, no studies on mass balances have been done so far. The latter is indispensable to understand transformation of phosphorus in RAS, an insight which is required to define the most optimal treatment method for removal or recovery of phosphorus from RAS.
Table 2-6 TP and RP in backwash waters of drumfilters of RAS
Type of RAS TP averaged
(mg-P/L) RP (mg-P/L) Reference
Fresh 12.5±6.3 a 4.0 ±1.9 a Ebeling and Ogden (2004) Fresh 77.8 ±7.8 a 4.0 ±1.9 a Ebeling et al. (2006) Fresh 59.3 ±11.5 a 22.8±6.57 a Rishel and Ebeling (2006) Saline 19.2-57.2 - Sharrer et al.(2007) Fresh 33.9-42.1 1.0-1.6 Sharrer et al.(2009) a standard deviation
Table 2-7 Phosphorus percentages in concentrated sludges from RAS Type of RAS (Fish species) TP in dry mass (%TS) Reference Brackish (Trout) 3.0-4.4 a Mudrak (1981) Fresh/Marine (-) 0.2-2.0 Chen et al. (1997) Fresh (Tilapia) 5.0 Shnel et al. (2002) Fresh (Eel) 12.0 Suzuki et al. (2003) Saline (Salmon) 1.6a Gebauer (2004) Fresh (Red tilapia) 3.9 Rafiee and Saad (2005) Saline (Salmon) 2.3 a Gebauer and Eikebrokk
(2006) Brackish (Striped
bass/tilapia) 1.1-3.1 Mirzoyan et al. (2008)
a
recalculated on the basis of TP and TS of sludge
The high phosphorus content of RAS sludge can be explained by several reasons. Whereas most RAS include biological nitrogen removal units like biofilters to eliminate accumulating nitrogen, no appropriate phosphorus removal and/or recuperation units are installed, meaning that phosphorus will accumulate (Martins et al., 2010). Martins et al. (2010) further state that the prevailing water management and legislation, resulting in the lack of specific phosphorus removal units, are also causing the discharge of P-rich effluents. Moreover, the efficacy and the cost-effectiveness of current phosphorus removal technologies are likely significant barriers to install phosphorus removal units in RAS.
On the basis of the high nitrogen and phosphorus contents of sludges from fresh RAS, their land application as fertilizer often has been practiced. However, land application of sludges from marine/brackish RAS seems to be unfeasible due to the high salts contents that might cause contamination of soils, groundwater and local water bodies via infiltration and/or run-off. In this aspect, application of AD to treat RAS sludges offers interesting possibilities to for concomitant bioenergy and phosphorus recovery from RAS sludges. In AD processes, reactive phosphorus is released from bound or particulate form of phosphorus, which provides possibilities to recuperate phosphorus from the solids fraction. Many studies and successful applications have been conducted in phosphorus recovery via struvite crystallization from digestates of AD
of sludges rich in phosphorus, such as waste activated sludges (Jaffer et al., 2002; Takiguchi et al., 2004), dairy cattle manure (Rico et al., 2011), piggery manure (Karakashev et al., 2008; Wrigley et al., 1992), and sludges from enhanced biological phosphorus reactors (EBPR) (Marti et al., 2008a; Marti et al., 2008b; Munch & Barr, 2001; Pastor et al., 2008). To our current insights and considering the cation composition of marine RAS, we feel that recuperation of phosphorus as struvite from the sludges from marine/brackish RAS has interesting potentials to reduce phosphorus emission in RAS. However, no reports have been found on this aspect, and hence further investigation is needed.
2.4 Prospects on AD in RAS
As aforementioned above, land-based tanks and enclosed bag net pens, for instance, particularly marine/brackish RAS, are a very promising approach to satisfy the increasing demand for seafood. However, in addition to the issues of the relatively concentrated waste stream produced and phosphorus accumulation in RAS, the lack of denitrification units and/or incomplete denitrification may also be problematic for the sustainability of RAS (Tal et al., 2009; Tal et al., 2006). Nitrate and nitrite produced from nitrification units can accumulate inside the major loop of RAS. Nevertheless, high concentrations of nitrate and particularly nitrite can present severe toxicity to fish (Menasveta et al., 2001). Thus, properly operated denitrification units are vital to RAS. However, generally external carbon sources are needed for efficient denitrification, which will increase the costs of fish farms. Thus, using AD to produce VFA from the relatively concentrated stream to supply carbon source to the denitrification is of great interest to be investigated for the future work.
In addition, it should be noted that anammox processes and the presence of ammonia oxidizing archaea (AOA) in saline conditions be responsible for at least 50% removal of nitrogen under marine conditions (Brown et al., 2013; Kartal et al., 2010). Therefore, investigation on the functionalities of anammox bacteria and AOA as well as those of ammonia oxidizing bacteria and nitrate oxidizing bacteria in brackish/marine RAS may be of great significance to improve denitrogenation efficiency. The application in RAS of cost effective autotrophic denitrification based on anammox and AOA allows a further decrease in energy costs and costs associated with external carbon sources for denitrification (Francis et al., 2007; Kartal et al., 2010; Tal et al., 2006).
2.5 Conclusions
The causes responsible for the low biogas production in the studies that have been conducted on sludge digestion in marine/brackish RAS are diverse, mainly including the use of improper inoculum and most importantly, high salinity, particularly high sodium concentration. The following topics require further research and will be conducive to improving biogas yields in the digestion of sludges from brackish or marine RAS and maintain its sustainability, meanwhile, recovering resources from RAS wastes such as phosphorus.
1) Use of saline and/or salinity adapted inoculum, e.g. marine sediments or sludges long-term adapted to environments with high salinity; analysis of the population evolution during inoculation and the long term run;
2) Further characterization of sludge from marine/brackish RAS, in terms of the contents of protein, lipid and carbohydrate;
3) Effective concentration methods of solids such as belt filter or coagulation/flocculation methodologies;
4) Use of RAS sludges for the anaerobic production of VFA to be subsequently used as carbon sources for denitrification units;
5) Impact of LCFA on AD of salty sludges;
6) Research on mass balance on phosphorus fractions in faeces, uneaten feed and bacterial biomass of sludges as well as on phosphorus recovery as struvite in the marine conditions;
7) Further investigation on the autotrophic denitrogenation systems based on anammox processes in marine/brackish RAS and on the profile of populations such as the potentially abundant AOA in marine/brackish RAS.
The authors believe that AD of the sludges from marine/brackish RAS offers a practicable approach to waste management of RAS, which aims not only to achieve a recuperation of both phosphorus and biogas and an effective sludge reduction, but also to cause less adverse impacts of RAS on local environments. In addition, phosphorus and nitrogen related issues are also of great interest and significance to be further examined in the brackish/marine RAS.
2.6 References
Abou-Elela, S.I., Kamel, M.M., Fawzy, M.E. 2010. Biological treatment of saline wastewater using a salt-tolerant microorganism. Desalination, 250(1), 1-5.
