toxicants inhibiting anaerobic digestion: a review
TRANSCRIPT
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Toxicants Inhibiting Anaerobic Digestion: A Review
Jian Lin Chen, Raphael Ortiz, Terry W.J. Steele, David C. Stuckey
PII: S0734-9750(14)00154-2DOI: doi: 10.1016/j.biotechadv.2014.10.005Reference: JBA 6849
To appear in: Biotechnology Advances
Received date: 28 March 2014Revised date: 8 October 2014Accepted date: 8 October 2014
Please cite this article as: Chen Jian Lin, Ortiz Raphael, Steele Terry W.J., StuckeyDavid C., Toxicants Inhibiting Anaerobic Digestion: A Review, Biotechnology Advances(2014), doi: 10.1016/j.biotechadv.2014.10.005
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Toxicants Inhibiting Anaerobic Digestion: A
Review
Jian Lin Chen a, Raphael Ortiz
b, Terry W.J. Steele
b *, and David C. Stuckey
a, c *
a Advanced Environmental Biotechnology Centre, Nanyang Environment & Water Research
Institute, Nanyang Technological University, Singapore 637141.
b School of Materials Science & Engineering, College of Engineering, Nanyang Technological
University, Singapore 637141.
c Department of Chemical Engineering, Imperial College London, London SW7 2AZ, and
Nanyang Environment & Water Research Institute, Advanced Environmental Biotechnology
Centre, Nanyang Technological University, Singapore 637141.
*Corresponding authors at: c
Department of Chemical Engineering, Imperial College
London, London SW7 2AZ, UK. Tel.: +44 207 5945591; fax: +32 15 317453. E-mail
address: [email protected] (David C. Stuckey); b
School of Materials Science &
Engineering, Nanyang Technological University N4.1-01-29, 50 Nanyang Drive,
Singapore 639798. Tel: +65-6592-7594, Fax: +65-6790-9081, Email address:
[email protected] (Terry W.J. Steele).
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Content
1 Introduction
2 Toxicants
2.1 Organic Toxicants
2.1.1 Chlorophenols
2.1.1.1 Toxicity of chlorophenols to anaerobic digestion
2.1.1.2 Mechanism of chlorophenols’ toxicity
2.1.2 Halogenated aliphatics
2.1.2.1 Toxicity of halogenated aliphatics to anaerobic digestion
2.1.2.2 Mechanism of halogenated aliphatics toxicity
2.1.3 Long chain fatty acids
2.1.3.1 Inhibition of long chain fatty acids to anaerobic digestion
2.1.3.2 Mechanism of long chain fatty acids inhibition
2.2 Inorganic Toxicants
2.2.1 Ammonia
2.2.1.1 Toxicity of ammonia to anaerobic digestion
2.2.1.2 Mechanism of ammonia toxicity
2.2.2 Sulfide
2.2.2.1 Toxicity/inhibition of sulfide to anaerobic digestion
2.2.2.2 Mechanisms of sulfide toxicity/inhibition
2.2.3 Heavy Metals
2.2.3.1 Toxicity of heavy metals to anaerobic digestion
2.2.3.2 Mechanisms of heavy metal toxicity
2.3 Nanomaterials
2.3.1 Toxicity/inhibition of nanoparticle/nanotubes to anaerobic
digestion
2.3.2 Mechanisms of toxicity/inhibition of nanoparticle/nanotubes
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3 Conclusions
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Abstract
Anaerobic digestion is increasingly being used to treat wastes from many sources because
of its manifold advantages over aerobic treatment, e.g. low sludge production and low
energy requirements. However, anaerobic digestion is sensitive to toxicants, and a wide
range of compounds can inhibit the process and cause upset or failure. Substantial
research has been carried out over the years to identify specific inhibitors/toxicants, and
their mechanism of toxicity in anaerobic digestion. In this review we present a detailed
and critical summary of research on the inhibition of anaerobic processes by specific
organic toxicants (e.g., chlorophenols, halogenated aliphatics and long chain fatty acids),
inorganic toxicants (e.g., ammonia, sulfide and heavy metals) and in particular,
nanomaterials, focusing on the mechanism of their inhibition/toxicity. A better
understanding of the fundamental mechanisms behind inhibition/toxicity will enhance the
wider application of anaerobic digestion.
Keywords: Anaerobic digestion; Inhibition; Toxicant; Chlorophenols, Halogenated
aliphatics, Long chain fatty acids, Nanomaterials, Ammonia, Sulfide, Heavy metals.
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List of Abbreviations
ADM1 Anaerobic Digestion Model No 1 AGS anaerobic granular sludge
CF chloroform CP monochlorophenol DCP dichlorophenol DCM dichloromethane DGGE denaturing gradient gel electrophoresis
DWCNT double walled carbon NTs
EGB expanded granular sludge bed EPS extracellular polymeric substances
FA free ammonia
FAN free ammonia nitrogen
IC50 half maximal inhibitory concentration
LCFA long chain fatty acid
MPB methane producing bacteria
NP nanoparticle
NT nanotube
NZVI nano zero valent iron
PCE perchlorethylene
PCP pentachlorophenol PEG polyethylene glycol chain
QSAR quantitative structure-activity relationship
ROS reactive oxygen species
SRB sulfate reducing bacteria
SWNT single-walled NT
TAN total ammonia nitrogen
TCE trichloroethylene TCP trichlorophenol
TeCP tetrachlorophenol TRFLP Terminal restriction fragment length
polymorphism UASB upflow anaerobic sludge bed
WWTP wastewater treatment plant
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1. Introduction
Anaerobic digestion can reduce organic pollution from the liquid outputs of homes,
industry and agriculture, while potentially offsetting the use of fossil fuels at the same
time. In addition, it offers numerous other significant advantages, such as lower energy
requirements, and less sludge production compared with traditional aerobic treatment
(Chen et al., 2008). Anaerobic digestion consists of a series of microbial processes that
convert organics to methane and carbon dioxide, and can take place under psychrophilic
(<20°C), mesophilic (25−40°C) or thermophilic (50−65°C) conditions, although
biodegradation under mesophilic conditions is most common. It also enables higher
loading rates than aerobic treatment, and a greater destruction of pathogens (Ravuri,
2013). Anaerobic digestion can be divided into three major microbial steps, i.e.
hydrolysis, acidogenesis/acetogenesis, and methanogenesis represented in Figure 1
(Amaya et al., 2013). During hydrolysis, a consortia of bacteria break down complex
organics (e.g. proteins, cellulose, lignin, and lipids) from the influent into soluble
monomers such as amino acids, simple sugars, glycerol, and fatty acids. Hydrolysis of
these complex polymers, some of which are insoluble, is catalyzed by extracellular
enzymes such as cellulases, proteases, and lipases (Batstone and Jensen, 2011).
Acidogenesis includes fermentation and anaerobic oxidation (β-oxidation), which are
carried out by fermentative acidogenic and acetogenic bacteria, respectively (Batstone
and Jensen, 2011). Fermentative acidogenic bacteria convert sugars, amino acids, and
fatty acids to organic acids (e.g. acetic, propionic, formic, lactic, butyric, or succinic
acids), alcohols and ketones (e.g. ethanol, methanol, glycerol, and acetone), acetate,
carbon dioxide, and hydrogen. Acetate is the major product of carbohydrate fermentation.
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Acetogenic bacteria convert fatty acids (e.g. long chain fatty acids) and alcohols into
acetate, hydrogen, and carbon dioxide, which are used by the methanogens. In the
methanogenesis step, acetate, hydrogen, and carbon dioxide are converted into methane
by methanogenic microorganisms, which are also classified as archaea composed of both
gram-positive and gram-negative bacteria with a wide variety of shapes, e.g., coccoid and
bacilli (Michael and Constantinos, 2006). Hydrolysis of insoluble polymers is generally
considered as rate-limiting among these successive steps, although with soluble feeds
methanogenesis is regarded as the key step in anaerobic digestion (Appels et al., 2008).
One of the main drawbacks to anaerobic digestion is its higher sensitivity to toxicants
than aerobic treatment. With the rapid development of nanotechnology, emerging
nanomaterials are starting to be used in some industrial products and will inevitably be
released into the environment; some nanomaterials have already been found in
wastewater treatment plants (WWTPs) and waste sludge (Yang et al., 2013). Hence, more
attention is being paid to their impact on the environment (Ju-Nam and Lead, 2008; Yang
et al., 2013), and in this review, we will summarize recent work on the effect of
nanoparticles and nanotubes on anaerobic digestion and the mechanisms by which they
may act. Besides, a wide range of organic chemicals can inhibit anaerobic digestion,
including halogenated benzenes (van Beelen and van Vlaardingen, 1994), halogenated
phenols (Armenante et al., 1999; Liu et al., 2008), phenol and alkyl phenols (Fedorak and
Hrudey, 1984; Levén et al., 2012), halogenated aliphatics (Adamson and Parkin, 1999;
Stuckey et al., 1980; van Hylckama Vlieg and Janssen, 2001), and long chain fatty acids
(LCFAs) (Hwu et al., 1996; Palatsi et al., 2012). In addition, many inorganic compounds,
such as ammonia (Ho and Ho, 2012; Liu and Sung, 2002), sulfide (Cai et al., 2008; Lopes
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and Lens, 2011) and heavy metals (Altaş, 2009; Oleszkiewicz and Sharma, 1990) are also
reported to be inhibitory, and the mechanisms behind inhibition are gradually becoming
understood. In this review we have decided to focus on the more common and typical
toxicants found in anaerobic membrane bioreactors (Casu et al., 2012; Feng et al., 2013;
Stuckey, 2012). This review will also present a detailed comparative summary of
research on the inhibition of anaerobic processes by nanomaterials, specific organic (i.e.,
chlorophenols, halogenated aliphatics and long chain fatty acids) and inorganic toxicants
(i.e., ammonia, sulfide and heavy metals), and critically analyze their mechanism of
inhibition.
2. Toxicants
2.1 Organic Toxicants
2.1.1 Chlorophenols
Chlorophenols are a group of chemicals produced by adding chlorine to phenol, and
include monochlorophenols (CPs), dichlorophenols (DCPs), trichlorophenols (TCP),
tetrachlorophenols (TeCPs), and pentachlorophenol (PCP). Chlorophenols are used
widely as pesticides, herbicides, antiseptics and fungicides as well as preservatives for
wood, glue, paint, vegetable fibers and leather. They are found to be highly persistent in
both aquatic and terrestrial environments, and are harmful to humans due to their
carcinogenicity (Muller and Caillard, 1986). Therefore, chlorophenols are listed as
priority pollutants by the U.S. Environmental Protection Agency (U.A. EPA).
2.1.1.1 Toxicity of chlorophenols to anaerobic digestion
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Although biodegradable, chlorophenols have been shown to be inhibitory to anaerobic
systems (Hernandez and Edyvean, 2008; Puyol et al., 2012; Wang et al., 1991). The
toxicity of chlorophenols depends on their degree of chlorination, the position of the
chlorine, and the purity of the sample; generally the level of inhibition increases with the
number of chlorine substitutions. For example, TCPs are all significantly more inhibitory
than DCPs, while PCP is considered to be the most toxic, e.g. 0.5-10 mg/L PCP can
cause inhibition to acidogens and methanogens (Chen et al., 2008). Adsorption tests
suggest that biological dechlorination is the main process for PCP removal in anaerobic
digestion, and due to the dechlorination high PCP removal efficiencies can be obtained
with the formation of DCP as the primary intermediate, followed by TCP and TeCP
(Damianovic et al., 2009). It was found that the partial dechlorination products of PCP,
such as the intermediate 3,4,5-TCP, is more toxic than PCP itself (Wu et al., 1989). For
chlorophenols with the same number of chlorine groups, the position of the substitution
influenced toxicity, and the sequence in decreasing order of toxicity was - meta para
ortho (Pera-Titus et al., 2004). By studying the biodegradation of chlorophenols in
anaerobic propionate-fed systems, Jin and Bhattacharya (1997) reported that the toxicity
of TCPs decreased in the following order: 2,4,5-TCP > 2,3,5-TCP > 2,4,6-TCP > 2,3,6-
TCP; while the six DCPs and three CPs showed slight toxicity to both propionate and
acetate degradation. In contrast, it has been reported that the chlorine position on CPs had
no significant effect on toxicity for either propionate or acetate degradation in sulfate
reducing anaerobic systems (Uberoi and Bhattacharya, 1997). Ennik-Maarsen et al.
(1998), however, found that during treatment of potato processing wastewater with a full
scale upflow anaerobic sludge blanket (UASB) reactor, the methanogenic activity of the
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sludge was inhibited more by 3-CP and 4-CP than by 2-CP; and this order can be related
to solvent/water partition coefficients (Kishino and Kobayashi, 1994) and these values
can be seen in Figure 2. In another study of chlorophenol’s toxicity to anaerobic sludge
from a sugar factory wastewater plant with high methanogenic activity, Vallecillo and
Vallecillo (1999) found that PCP showed the highest toxicity, and they concluded that
chlorophenol toxicity increased with the number of substituted chlorine atoms, except
2,4-DCP which was more toxic than 2,4,6-TCP. The toxicity of various chlorophenols
towards the syntrophic methanogenic reaction is summarized in Figure 2.
By now it is well known that substrate structure and concentration have an important
influence on methanogenic inhibition; however, the initial biomass concentration has not
always been taken into account. The initial biomass concentration is important since it
defines the initial substrate to microorganism ratio, S0/X0. This ratio determines whether,
in a given batch system, the cell will multiply or only grow without dividing (Chudoba et
al., 1991). For example, when studying the influence of S0/X0 on the inhibition of
methane production caused by 4-chlorophenol (4-CP), it was observed that inhibition
decreases as S0/X0 decreases (initial biomass concentration increases), and that for the
same value of S0/X0 there was an increase in inhibition when X0 decreases (Moreno
Andrade, 2003). In addition, based on the assumption that biomass growth is negligible
with respect to methane production, i.e. biomass concentration can be considered
constant during the biological tests, Puyol et al. (2012) proposed a kinetic model to
predict the inhibition of methanogenesis caused by chlorophenols and recommended it as
a useful tool to improve the recognized methanogenesis modeling by the IWA Anaerobic
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Digestion Model No 1 (ADM1) (Batstone et al., 2002) when treating inhibitory and/or
toxic compounds with anaerobic technologies.
2.1.1.2 Mechanism of chlorophenol toxicity
It is widely believed that only molecules of cyclic hydrocarbons dissolved in the aqueous
phase are available for intracellular metabolism (Wodzinsk.Rs and Bertolin.D, 1972).
Transfer of hydrophobic substrates, such as chlorophenols, proceeds via dissolving in the
aqueous phase and subsequently being taken up by the cell; however, direct contact
between the cell membrane’s hydrophobic components and hydrophobic compounds is
prevented by the cell wall and/or particularly, the cell membrane’s hydrophobic
components (Sikkema, 1995). Partitioning of these hydrophobic molecules into the lipid
bilayer of the cytoplasmic membrane is the most important mechanism in the uptake of
hydrophobic compounds, and may result in the accumulation of these compounds in the
lipid bilayer to enhance their availability to the cell and may cause toxicity problems as
well (Sikkema et al., 1994). Partition coefficients of hydrophobic compounds in
octanol/water, olive oil/water, diethyl ether/water, hexadecane/water, etc. have been used
to predict the effects of these compounds on intact cells (e.g., bioconcentration,
biodegradation and toxicity) (Osborne et al., 1990). Sierra-Alvarez and Lettinga (1991)
found that the logarithm of the partition coefficient in octanol/water (logP), an indicator
of hydrophobicity, was positively correlated with methanogenic inhibition, suggesting
that hydrophobicity is an important factor contributing to the toxicity of most inhibitory
aromatic compounds. Ennik-Maarsen et al. (1998) also obtained a high correlation
between logP for chloro-substituted benzenes and phenols and their methanogenic
toxicity. In addition, it has been proved that the protein-to-lipid ratios in the membrane
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and membrane functioning could be significantly influenced by phenols (Keweloh et al.,
1990). Finally, Heipieper et al. (1991) reported that addition of phenol and 4-
chlorophenol to suspensions of E. coli induced an efflux of potassium ions. Nevertheless,
the inhibitory action of chlorophenols seems to be directly related to these lipophilic
compounds’ partitioning behavior, which is expressed as logP. This suggests that the
(cytoplasmic) membrane is the primary site for the toxic action of chlorophenols by
disrupting the proton gradient through the membrane, and interfering with cellular energy
transduction, thereby decreasing cell growth due to an uncoupling of the catabolic and
anabolic reactions (Chen et al., 2008).
2.1.2 Halogenated aliphatics
Halogenated aliphatics (HAs) are organic chemicals in which one or more hydrogen
atoms has been replaced by a halogen. HAs are used in industry as solvents, chemical
intermediates, and fumigants and insecticides, and are found in the chemical, paint and
varnish, textile, rubber, plastics, dye-stuff, pharmaceutical and dry-cleaning industries
(Fishbein, 1986). Many of these synthetic compounds, especially the chlorinated
insecticides, due to their poor bioavailability, xenobiotic structure or high toxicity of the
compound itself or of intermediates, are recalcitrant and persist in the environment,
although some can be degraded under certain conditions (Cappelletti et al., 2012).
2.1.2.1 Toxicity of halogenated aliphatics to anaerobic digestion
Although microorganisms have been reported with the ability to remove halogens from
aliphatic compounds by the activity of enzymes known as dehalogenases (Slater, 1994),
most of the HAs are strong methanogenic inhibitors (Yu and Smith, 2000). Among HAs
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it has been suggested that brominated aliphatics, in general, are more inhibitory to
methanogens than their chlorinated analogs (Belay and Daniels, 1987; Freedman, 1991).
Among the most frequently encountered contaminating HAs in the environment are
polychlorinated aliphatic hydrocarbons, such as dichloromethane (DCM, CH2Cl2),
chloroform (CF, CHCl3), trichloroethylene (TCE, C2HCl3), and perchlorethylene (PCE,
C2Cl4), which are common priority pollutants found in wastewaters, soils, and aquifers
(Riley and Zachara, 1992).
Chloroform (CF) is the most widely used chloroaliphatic whose methanogenic toxicity
has been studied widely. By studying the inhibitory effect of DCM, CF, TCE, and PCE,
Yu and Smith (2000) found that CF was the most inhibitory, and at an aqueous
concentration at 0.09 mg/L CF inhibited methanogenesis completely; while DCM
inhibited methanogenesis at 3.9 mg/L, TCE at 18 mg/L, but PCE did not inhibit
methanogenesis at 14.5 mg/L. Furthermore, among six chlorinated hydrocarbon solvents
including 1,1,1-trichloroethane and carbon tetrachloride, CF was reported to be the most
toxic to the anaerobic digestion of sewage sludge (Swanwick and Foulkes, 1971).
Meanwhile, with the concentration inhibiting the mineralisation rate by 10% (IC10) at
0.04 mg/L, CF proved to be extremely toxic to anaerobic mineralisation of 4-
chlorophenol in microcosms with methanogenic sediment compared to benzene (IC10 at
150 mg/L), 1,2-dichloroethane (IC10 at 71 mg/L) and pentachlorophenol (6 mg/L) (van
Beelen and van Vlaardingen, 1994). The toxicity of CF to many obligate anaerobic
prokaryotes has been reported (Weathers and Parkin, 2000). One mg/L of CF can
completely inhibit dechlorination of PCE by a chlororespiring anaerobic isolate (Maymo-
Gatell et al., 2001), suggesting inhibition by CF for reductive dechlorination of
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chloroethenes is a general problem for sites co-contaminated with CF, which can only be
overcome by first removing the CF (Bagley et al., 2000). However, it has been suggested
that the high toxicity of CF to methanogenesis might be attributed to the formation of
very toxic and reactive intermediates during the slow anaerobic degradation of CF in
anaerobic sludge, although methanogenic toxicity data for these intermediates are lacking
(Trohalaki et al., 2003). Terminal restriction fragment length polymorphism (TRFLP)
analyses combined with a clone library showed that CF inhibited not only methanogenic
activity, but also the structure of methanogenic communities (Xu et al., 2010) because the
acetoclastic Methanosaetaceae were more sensitive to CF than hydrogenotrophic
Methanobacteriales and Methanomicrobiales. Xu et al. (2010) suggested that it is
probably some hydrogenotrophic methanogens, such as Methanobacteriaceae and
Methanomicrobiaceae, which can synthesize coenzyme M (CoM) which exhibit lower
rates of transport of external CoM into the cell, and are likely to be more resistant to CF
and other toxicants such as 2-bromoethanesulfonate.
2.1.2.2 Mechanism of halogenated aliphatics toxicity
Unlike chlorophenols, there is no identified relationship between the number of chloro-
substituents and the toxicity of chloroaliphatics. Wimley and White (1993) proposed that
polarity is an important factor, however, this is insufficient for predicting the toxicity of
different HAs. Furthermore, Gill and Ratledge (1972) demonstrated that, besides
halogenated substituents, the toxicity of HAs is related to their chain length, perfectly
correlating with their hydrophobicity and their solubility in water. Originally developed
for pharmaceutical applications (Leo, 1975), quantitative structure-activity relationships
(QSARs) have also been successfully used to predict environmental toxicity (Erturk et al.,
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2012; Jensen et al., 2008; Wu et al., 2013). Although QSARs have been helpful in
assessing the toxicity of specific groups of chemicals (e.g., phenols, chlorinated
aromatics), and also have been applied to toxicity prediction for HAs (Trohalaki and
Pachter, 2003; Trohalaki et al., 2003), they do not provide any information on the
mechanisms of these toxic effects, and this needs further investigation. By coordinating
the study on inhibition of methanogenesis by DCM, CF, TCE, and PCE, Yu and Smith
(2000) proposed a model to explain why corrinoids and porphinoids are not only the
dechlorination catalysts, but also the target moieties by which CF inhibits methanogens
(Figure 3). This model proposes that CF can be either a direct or indirect methanogenic
inhibitor due to its molecular structure, which is similar to some of the key C-1
intermediates of the methanogenesis pathway. In direct inhibition, due to its high redox
potential (10.56 V), CF would bind to the intracellular reduced corrinoid and porphinoid
enzymes (pool B, Figure 3) of the methanogenesis pathway, resulting in a higher affinity
for these reduced enzymes. This would prevent the normal substrate binding, such as a
methyl group, and channel electron flow away from methanogenesis towards
dechlorination (Figure 3), and hence CF’s binding to these enzymes leads to direct
inhibition of methanogenesis. Normal methanogenesis will resume when the
corrinoid/porphinoid enzymes are freed after CF is depleted. This model explains how
CF is dechlorinated reductively, and why the inhibition of methanogenesis is alleviated
once CF is dechlorinated. On the other hand, the binding of CF to the intracellular free
corrinoids/porphinoids lowers the free corrinoid/porphinoid concentrations (pool F, Fig.3)
in the cell, and could shift the equilibrium between the intracellular pools of free
corrinoids/porphinoids and protein-bound corrinoids/porphinoids (pool B, Fig.3) to the
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free-form side. This would result in the indirect inhibition of methanogenesis because
protein-bound corrinoids and porphinoids are essential for methanogenesis. In addition,
electron sinks of intermediates of CF and free corrinoids/porphinoids (pool F, Fig.3)
would drain the electron flow from normal methanogenesis to CF dechlorination, and this
in turn would indirectly inhibit methanogenesis.
2.1.3 Long chain fatty acids
Fatty acids are chains of carbons with attached hydrogen molecules at one end, and an
acid group at the other end. Long chain fatty acids (LCFAs) are fats that have several
carbons in their chain and include unsaturated and saturated fats. LCFAs are important
fractions of the organic matter in oil/fat wastewater (Stoll and Gupta, 1997). Anaerobic
digestion has been used for many decades to treat oily/fatty wastes, e.g., ice-cream wastes
(Hawkes et al., 1995), dairy wastewater (Perle et al., 1995), fish waste (Achour et al.,
2000), slaughterhouse wastewater (Cuetos et al., 2008) and vegetable waste (Li et al.,
2002) from the food processing industries. These lipids are neither easily treated by
conventional means, nor decomposed biologically, due to their formation of insoluble
aggregates and floating on the surface of the wastewater (Stoll and Gupta, 1997).
2.1.3.1 Inhibition of long chain fatty acids to anaerobic digestion
Although lipid rich wastewaters have high methane potential, one of its intermediate
products, LCFAs, can lead to inhibition (Palatsi et al., 2009). Hanaki et al. (1981) found
that the addition of LCFAs caused the appearance of a lag period in methane production
from acetate, and in the anaerobic digestion process. According to the Second Law of
Thermodynamics, energy must flow from a higher to a lower potential.
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Thermodynamically acido-acetogenesis (fermentation) is endothermic and non-
spontaneous (ΔH > 0 and ΔG > 0), and this implies that LCFA fermentation cannot
proceed spontaneously, and that LCFAs form insoluble hydrophobic aggregates in the
aqueous phase (Oh and Martin, 2010). However, acidogenesis together with acetogenesis
decomposes the long chain saturated fatty acids to acetic acid though shorter chain fatty
acids (β-oxidation) (Oh and Martin, 2010) and by electrochemical coupling with
methanogenesis, a strongly spontaneous process with a negative enthalpic driving force
(ΔG ≪ 0 and ΔH < 0) (Oh and Martin, 2007) occurs, and the non-spontaneous
fermentation is thermodynamically switched to a spontaneous process (ΔG < 0 and ΔH <
0) (Bartoschek et al., 2000). This implies that LCFA inhibition of methanogenesis could
cause the failure of LCFA fermentation and the whole anaerobic digestion bioprocess.
Severe LCFA inhibition was observed in an anaerobic membrane bioreactor for the
treatment of lipid rich corn-to-ethanol thin stillage, and Dereli et al. (2014) suggested that
the extensive dissolution of LCFAs in the reactor broth possibly caused inhibition of
methanogenesis. Inhibition of methanogens by LCFAs is reported as the limitation to
exploit biogas production from fat-rich wastewaters (Silva et al., 2014). Inhibition by
LCFAs of anaerobic processes depends on the type of LCFA, the microbial population,
and the temperature (Silvestre et al., 2014). For instance, oleic acid, followed by palmitic
and stearic acids, has been described as the LCFA with the greatest inhibitory effect on
thermophiles (Pereira et al., 2005).
2.1.3.2 Mechanism of long chain fatty acid inhibition
Although LCFA degradation was suggested as the “limiting step” in methanogenesis due
to its perceived limitations (Novak and Carlson, 1970), by combining hydrolytic,
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fermentative, syntrophic acetogenic (SAB) and methanogenic microorganisms, anaerobic
digestion of LCFAs has been achieved (Summers and Bousfield, 1980). The limiting step
was suggested to be closely related to the initial concentration of LCFAs, and therefore
high concentrations of LCFAs lead to the failure of anaerobic digesters (Chen and
Hashimoto, 1978; Tijero et al., 1989). It is believed that inhibition of anaerobic
metabolism by LCFAs is because of the adsorption of the LCFAs onto the cell wall and
membrane affecting metabolic transport (Alves et al., 2001; Masse et al., 2002; Neves et
al., 2006). This adsorption delays methane production, but can be prevented by providing
a competitive synthetic adsorbent (such as bentonite) (Figure 4) (Palatsi et al., 2009), and
the bio-physics of LCFA inhibition in anaerobic digestion has been mathematically
modeled by Zonta et al. (2013). In addition, Hwu et al. (1996) proved that the toxicity of
a model LCFA (oleate) to acetoclastic methanogens in anaerobic sludge was not
dependent on three biological factors (sludge origin, specific acetoclastic methanogenic
activity and sludge adaptation to lipids), but was closely correlated to the physical factor
of specific surface area of the sludge. Besides the inhibition of methanogenic bacteria,
Pereira et al. (2005) and Neves et al. (2009) proposed that LCFA adsorption and
accumulation on biomass can create a physical barrier and hinder the transfer of
substrates and products, inducing an initial delay in methane production, and even
causing sludge flotation and washout. Addition of calcium has been shown to reduce
LCFA inhibition, probably due to the formation of insoluble salts (Hanaki et al., 1981).
Dilution of the reactor’s content with inoculum, thus increasing the biomass/LCFA ratio,
or the addition of adsorbents, were found to be the best strategies to recover thermophilic
manure reactors subjected to LCFA inhibition (Palatsi et al., 2009).
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2.2 Inorganic Toxicants
2.2.1 Ammonia
2.2.1.1 Toxicity of ammonia to anaerobic digestion
Ammonia is an essential nutrient for the growth of microorganisms involved in anaerobic
digestion, as well as acting as an inhibitor at certain high concentrations (Koster and
Lettinga, 1984). The fermentation of nitrogen-containing materials such as urea and
proteins releases ammonia-nitrogen which exists largely as the ionized form (NH4+), but
this depends strongly on pH as its pKa is 9.3, and hence the toxic unionized form (NH3)
increases with increasing pH. The anaerobic digestion of livestock wastes releases high
levels of ammonia which raises the pH and forms higher levels of free ammonia (FA)
which is widely known to inhibit methanogenic microorganisms with low methane
production (Jin et al., 2012). Zhou and Qiu (2006) proved that concentrations of ammonia
nitrogen at 2.48 and 2.89 g/L can inhibit 50% of specific methanogenic activity in upflow
anaerobic sludge bed (UASBs) and expanded granular sludge bed (EGBs) reactors,
respectively. In terms of a 50% reduction in methane production, a wide range of
ammonia concentrations has been documented, with the inhibitory total ammonia
nitrogen concentration ranging from 1.7 to 14 NH3–N g/L (Chen et al., 2008). Some other
findings for the impact of ammonia on anaerobic digestion have been summarized in
Table 1. In a recent study on the impact of ammonia on thermophilic anaerobic digestion,
it was found that it was clearly affected by increasing concentrations of ammonia; the
threshold C/N ratio was found to be 4.40, corresponding to 620 mg free ammonia/L
(Siles et al., 2010). Calli et al. (2005) proved that acetogenic bacteria are more sensitive
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than methanogenic archaea to free ammonia, which has been suggested to be the active
component causing ammonia inhibition (Angelidaki and Ahring, 1993; He et al., 2011).
Meanwhile acetate-utilizing bacteria adapted to ammonia were shown to grow with a free
ammonia concentration of up to 800 mg-N/L, while many lower free ammonia
concentrations (100–150 mg-N/L) have been reported to initially inhibit an unadapted
process (Braun et al., 1981; De Baere et al., 1984).
The free ammonia concentration (CNH3) can be calculated according to the following
equation:
(1)
where TAN is the total ammonia nitrogen concentration, mg/L, Ka is the temperature
dependent dissociation constant (0.564 10-9
at 25 ºC, 1.097 10-9
at 35 ºC and 3.77
10-9
at 55 ºC), and CH is the concentration of hydrogen ions (Kayhanian, 1999).
Accordingly, the free ammonia concentration (CNH3) depends primarily on three
parameters; the total ammonia nitrogen concentration, temperature and pH (CH).
Ammonia concentrations below 200 mg/L are suggested to be beneficial to anaerobic
processes as this provides an essential nitrogen nutrient for anaerobic microorganisms
(Liu and Sung, 2002). Several studies found that the fermentation of high ammonia-
containing wastes is more easily inhibited at thermophilic temperatures than at
mesophilic temperatures (Angelidaki and Ahring, 1994; Bayr et al., 2012; Braun et al.,
1981), which is in agreement with the fact that the ratio of free ammonia to the total
ammonium will be much higher at higher temperatures as noted by Garcia and Angenent
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(2009). Therefore, wastewater rich in ammonium, urea and protein is more difficult to
treat under thermophilic conditions (i.e., 55–65 °C), even though the kinetics are more
favorable compared to mesophilic conditions (i.e., 25–37 °C) (Bocher et al., 2008; El-
Mashad et al., 2004). In aqueous solutions, there is a chemical balance between
ammonium ions (NH4+) and FA (NH3):
(2)
where obviously pH can affect the equilibrium of NH3/NH4+, and a high pH is conducive
to the formation of FA (Mosquera-Corral et al., 2005). Furthermore, Koster (1986)
reported that when the pH value increases, the biogas process becomes more sensitive
towards ammonia; from equation 2 an increase in pH results in more free ammonia. An
increase in pH from 7 to 8 will actually lead to an eight-fold increase in the free ammonia
concentration.
2.2.1.2 Mechanism of ammonia toxicity
Free ammonia is more toxic to methanogens than ionised ammonium (NH4+) because it
diffuses more rapidly through the cell membrane, causing proton imbalance, and/or
potassium (K+) deficiency, while ionised ammonium may just inhibit the methane
synthesizing enzyme directly (Gerardi, 2006; Kayhanian, 1999). When free ammonia
diffuses passively into methanogens, the difference in intracellular pH causes some of
them to convert to ammonium (NH4+), absorbing protons (H
+) in the process, while by
using a potassium antiporter, the cells then expend energy on proton balancing (Sprott et
al., 1984). However, the diffusion of free ammonia molecules into the cell also depends
on the physiology of the methanogens, e.g., on the basis of kilograms of NH3 entering per
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kilogram cell mass per hour, less free ammonia diffused into the cell of Methanosarcina,
which consist of larger spherical cells than into the smaller rod-shaped cell of
Methanothrix (Wiegant and Zeeman, 1986). Another reason why the ionized form of
ammonia is more beneficial than the free form is that the hydroxide, produced from
equation 2, can react with carbon dioxide, produced during the anaerobic digestion
process, to form bicarbonate (Kayhanian, 1999):
(3)
This will increase the buffering capacity of the anaerobic reactor, making the process less
susceptible to minor fluctuations in the relative production rates of the acetogenic and
methanogenic bacteria and, therefore, more stable.
A common approach to ammonia inhibition relies on dilution, and various types of
inhibition can be counteracted by increasing the biomass retention in the reactor.
Moreover, to mitigate the inhibition of thermophilic anaerobic treatment of digested
piggery wastewater by ammonia, Ho and Ho (2012) studied the effect of pH reduction,
zeolite, biomass and humic acid on methane production, and suggested that a reduction in
the pH from 8.3 to 6.5, and addition of 10–20 g/L zeolite, as more effective strategies
than addition of biomass and humic acid to mitigate ammonia inhibition.
2.2.2 Sulfide
Sulfide containing waste streams are generated by a number of industries such as
petrochemical plants, tanneries, viscose rayon factories, and coal gasification for
electricity production, or the anaerobic treatment of sulfate containing wastewater (Cai et
al., 2008).
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2.2.2.1 Toxicity/inhibition of sulfide to anaerobic digestion
Sulfidogens or sulfate reducing bacteria (SRB), which reduce sulfate to sulfide, play a
significant role in the anaerobic digestion of complex substrates. McCartney and
Oleszkiewicz (1991) found that SRB: i) generate sulfides which may be inhibitory and/or
toxic to SRB and methane producing bacteria (MPB); ii) reduce the rate of
methanogenesis; and iii) decrease the quantity of methane produced by competing for the
available carbon and/or hydrogen. In addition, Karhadkar et al. (1987) proposed two
stages of methanogenic inhibition due to sulfate reduction, viz. primary inhibition due to
competition for substrate from the SRBs, and secondary inhibition resulting from
methanogenic population decline due to inhibition of cellular function by soluble sulfides.
During the anaerobic treatment of wastewater containing sulfate, the competition
between SRB and MPB for acetate as their common primary substrate can affect
treatment efficiency. Although the primary competitive inhibition can be overcome in the
anaerobic treatment of sulfate-rich wastewaters, to maintain a low oxidation–reduction
potential in the reactors makes the presence of small amounts of sulfide advantageous. It
has been shown that sulfide ions can inhibit methane formation, suggesting non-
competitive inhibition of methanogenesis due to the sulfide resulting from SRB activity
may result in process failure (Paula Jr. and Foresti, 2009). It has been reported that high
levels of sulfide (IC50 ∼1300 mg/L H2S) did not affect SRB growing on acetate and
ethanol (Isa et al., 1986), and it has even been observed that sulfate removal rates
increased on ethanol and sugar with increasing total sulfide concentrations up to 1424
mg/L (Greben et al., 2005). However, sulfide was reported to be toxic to unacclimated
methanogens at concentrations of 50 only mg/L (Parkin et al., 1983), while
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Maillacheruvu and Parkin (1996) found that H2S inhibited acetotrophic methanogens
much more than hydrogenotrophic methanogens. Yamaguchi et al. (1999) reported that
the IC50 of H2S for acetotrophic and hydrogenotrophic methanogens was 160 and 220
mg/L, respectively.
2.2.3.2 Mechanisms of sulfide toxicity/inhibition
Theoretically, the reduction of sulfate to sulfide:
(∆G = −154 kJ) (4)
(∆G = − 43 kJ) (5)
yields more energy than methanogenesis:
(∆G = −135 kJ) (6)
(∆G = −28.5 kJ) (7)
which makes the latter noncompetitive and may reduce the rate of methanogenesis and
decrease the quantity of methane production (Karhadkar et al., 1987). In addition, sulfide
can denature proteins due to the formation of cross-links among polypeptide chains, and
can interfere with key metabolic enzymes in the cells (Madigan et al., 2003). Sulfide can
also interfere with the assimilation of sulfur and affect the intracellular pH (Visser, 1995).
Therefore, uncoupling growth from energy production and cell maintenance needs more
energy (Okabe et al., 1995). The toxicity of sulfide is often associated with its
undissociated form (H2S) due to the facilitated passage of neutral molecules across cell
membranes, and its high reactivity with cellular components (O'Flaherty et al., 1998).
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The chemical equilibrium of the sulfide species is pH dependent, with most of the total
sulfide (TS) in the HS− form at pH 8, while most is H2S at pH 6. The aqueous H2S
concentration can be calculated from the equilibrium equation (McCartney and
Oleszkiewicz, 1991):
(8)
Sulfide toxicity to methanogens is proportional to its concentration (on an added basis) in
the substrate and H2S concentration in the gas phase, which also incorporates the effect of
pH. Therefore, diluting the wastewater stream can prevent toxicity, although in general
this approach is considered undesirable due to the increase in the total volume of
wastewater that must be treated. Chen et al. (2008) suggested incorporating a sulfide
removal step in the overall process to reduce the sulfide concentration in an anaerobic
treatment system. In addition, adaptation of the MPB to free H2S, particularly in reactors
with biomass that cannot be washed out easily, eg biofilms and membrane systems, could
increase the tolerance of MPB to sulfide.
2.2.3 Heavy Metals
Heavy metals are often present in industrial wastewaters and municipal sludge in
significant concentrations, and the most frequently found are copper (Cu), zinc (Zn), lead
(Pb), mercury (Hg), chromium (Cr), cadmium (Cd), iron (Fe), nickel (Ni), cobalt (Co)
and molybdenum (Mo) (Altaş, 2009). However, many metals are required for the
activation or functioning of many enzymes and coenzymes in anaerobic digestion. Some
heavy metals, such as Ni, Co and Mo, are required at low concentrations, while the order
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of heavy metal composition in cells was found to be Fe >> Zn Ni > Co = Mo > Cu
when analyzing ten methanogenic strains (Takashima and Speece, 1989). However,
excessive amounts of heavy metals can lead to inhibition or toxicity (Li and Fang, 2007).
2.2.3.1 Toxicity of heavy metals to anaerobic digestion
The values of the half maximal inhibitory concentration (IC50) of different heavy metals
in other anaerobic digestion systems are summarized in Table 2. Unlike many other toxic
substances discussed above, heavy metals are non-biodegradable and can accumulate to
potentially toxic concentrations (Nayono, 2009; Sterritt and Lester, 1980). The potential
toxicity of heavy metals is controlled, to a large extent, by the physical and chemical
nature of the environment in which they are present, and this is correlated to different
ion-specific physicochemical parameters, e.g., standard reduction-oxidation potential,
electronegativity, the solubility product of the corresponding metal-sulfide complex, the
Pearson softness index, electron density and the covalent index (Workentine et al., 2008).
Heavy metals have been reported to be inhibitory to anaerobic microorganisms including
acetogens (Li and Fang, 2007), acidogens (Yu and Fang, 2001a; b; Zayed and Winter,
2000), methanogens (Karri et al., 2006; Mori et al., 2000) and SRB (Utgikar et al., 2003).
Evaluating the toxicity of heavy metals during the anaerobic digestion of sewage sludge
indicated severe inhibition at different concentration ranges for certain heavy metals,
such as from 70 to 400 mg/L for Cu, 200 to 600 mg/L for Zn and 10 to 2000 mg/L for Ni
(Ahring and Westermann, 1985). Moreover, lower concentrations of heavy metals
showing toxicity to anaerobic digestion were found in experiments under more defined
conditions. For example, the IC50 of acetate-degrading methanogens was reported to be
7.7, 12.5, 16 and 67.2 mg/L for Cd, Cu, Zn and Pb, respectively (Lin, 1992), and Cu, Zn
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and Ni caused 50% inhibition of mixed methanogens at 10, 40 and 60 mg/L, respectively
(Zayed and Winter, 2000). Acidogens are believed to be more resistant to heavy metal
toxicity than methanogens (Zayed and Winter, 2000). However, Hickey et al. (1989)
proposed that, compared to methanogenic populations, some trophic groups or organisms
within the anaerobic consortia might be more severely inhibited by a pulsed addition of
heavy metals. Studies have been carried out on the effect of heavy metals on a variety of
microbial species in anaerobic digestion. For example, with the addition of 5 mg Cd/L the
activity of Betaproteobacteria, which are involved in nutrient removal in activated sludge,
decreased significantly from 30.7% to 2.1% in an anaerobic–anoxic–oxic system
compared to the same system without the addition of cadmium (Tsai et al., 2005). In
another study on 2-CP degradation in anaerobic bioreactors, it was found the abundance
of the predominant archaeal species (e.g. Methanothrix soehngenii, Methanosaeta concilü
and uncultured euryarchaeota) decreased with Cd2+
and Cu2+
addition at high
concentration (shocked by 300 mg/L for 3 days), and the 2-CP anaerobic degradation
system was more sensitive to Cu2+
than Cd2+
(Huang, 2008). This observation is useful
with reference to the fundamentals, and the monitoring and control of anaerobic
membrane reactor responses to ramp/shock heavy metals loads.
2.2.3.2 Mechanisms of heavy metal toxicity
In an anaerobic environment, heavy metals may be: i) precipitated as sulfides (except Cr)
carbonates or hydroxides (Gonçalves et al., 2007; Gould and Genetelli, 1978); ii)
chelated and maintained in solution by compounds produced during digestion (Callander
and Barford, 1983; Walker et al., 2003), e.g. Soluble Microbial Products (SMPs); and iii)
adsorbed/bound to sludge ligands (Alibhai et al., 1985). At present it is believed that the
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failure of anaerobic digestion due to heavy metals occurs only when the concentration of
their free ions exceeds some threshold value (Leighton and Forster, 1997; Oleszkiewicz
and Sharma, 1990). Aluminum appears to be an exception as it becomes gradually
inhibitory only beyond 1500 mg/L as A12(SO4)3, and this high toxicity threshold could be
attributed to its low solubility under anaerobic conditions (Leighton and Forster, 1997). It
is believed that heavy metals show their toxicity due to their disruption of enzyme
function and structure by binding with thiol and other groups on protein molecules, or by
replacing natural metals in enzyme prosthetic groups (Soldatkin et al., 2012). In addition,
Oleszkiewicz and Sharma (1990) summarized various mechanisms of metal
toxicity/inhibition as: i) substitution of metallic enzyme cofactors (metal prosthetic
group); ii) combining with the outstanding sulfhydryl group (-SH) such as in cysteine; iii)
inactivation of the mercapto group in coenzyme M (2-mercaptoethanesulfonate) in
methanogens; and iv) tight binding to acid groups in the side chains of the amino acids in
the polypeptide chain.
Because heavy metals are usually present as mixtures in the influent of an anaerobic
digester, specific antagonistic and synergistic effects have been reported and are of great
interest in managing and ameliorating toxicity. Synergism means the enhanced toxic
effect of one metal in the presence of small amounts of another metal, while antagonism
refers to the effect of one metal which alleviates the toxic effect of another metal.
Oleszkiewicz and Sharma (1990) reviewed these phenomena in an early study on the
toxicity of Pb, Cu, Cd, Zn and Hg on anaerobic digestion, and showed that aluminum
silicate was antagonistic to the toxicity of other heavy metals. Another interesting heavy
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metal, nickel, has been proved to be synergistic in Ni-Cu, Ni-Mo-Co, and Ni-Hg systems,
and antagonistic in Ni-Hg, Ni-Zn, and Ni-Cd systems (Babich and Stotzky, 1983).
Heavy metal ions can be removed, or their concentrations reduced, by several
mechanisms such as precipitation, sorption and chelation by organic and inorganic
ligands, which are proposed as the most important methods for mitigating heavy metal
toxicity (Agrawal et al., 2011). For example, the addition of sulfide has been the main
method to precipitate heavy metals in anaerobic treatment (Kieu et al., 2011). Meanwhile,
Aquino and Stuckey (2007) proved that some insoluble sulfide salts were biologically
available under anaerobic condition. In addition, sorption of heavy metals onto activated
carbon, kaolin, bentonite, diatomite and waste materials such as compost and cellulose
pulp waste can also mitigate inhibition (Ulmanu et al., 2003). More recently, anaerobic
biological processes relying on the activity of SRB are being considered for the treatment
of heavy metal containing effluents (Kieu et al., 2011).
2.3 Nanomaterials
As the technological benefits of nanotechnology begin to move rapidly from laboratory to
large-scale industrial application (Abhilash, 2010; Schmid and Riediker, 2008), release of
nanomaterials to the environment is inevitable, and major environmental receptors will be
soil, sediment, and biosolids from wastewater treatment (Brar et al., 2010; Eduok et al.,
2013). Understanding the effect of nanoparticles on anaerobic microbial activity and
communities is important in order to enhance anaerobic process design.
2.3.1 Toxicity/inhibition of nanoparticle/nanotubes to anaerobic digestion
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The applications for nanoparticles (NPs, particles of any shape with dimensions in the
1×10−9
and 1×10−7
m range (Alemán et al., 2007)) are growing rapidly and outpacing
scientific investigations into the potential environmental impact as emerging contaminant.
The impact of metallic and metal oxide NPs (e.g. silver nanoparticles (Ag-NPs), nano-
ZnO, nano-TiO2, nano-Al2O3, nano-SiO2, nano-Au, nano-CeO2 and nano zero valent iron
(NZVI)) on sludge anaerobic digestion has been reviewed, suggesting that Ag-NPs, nano-
Al2O3, nano-SiO2 and nano-TiO2 are chemically stable NPs that have no adverse effects
on microbes under anaerobic conditions (Stasinakis, 2012; Yang et al., 2013), while
nano-Au presented no or low toxicity to anaerobic biomass and nano-CeO2 was the most
toxic to both mesophilic and thermophilic biomass (Stasinakis, 2012). When
investigating the behavior of Ag-NPs in a non-aerated tank, and in a pilot WWTP, Kaegi
et al. (2011) found that Ag-NPs transformed to Ag2S in less than 2 hours under anaerobic
conditions, and most of the Ag in both the sludge and effluent of the WWTP was present
in the form of Ag2S. In another batch anaerobic digestion system over less than two
weeks, Jin et al. (2012) observed that there was no significant difference in biogas or
methane production between the sludge treated with 40 mg Ag-NPs and the control. In
the same study, using quantitative PCR assays at moderate concentrations (≤ 40 mg/L),
Jin et al. (2012) found that Ag-NPs had no significant impact on methanogenic
assemblages and anaerobic production due to almost no release of Ag+ ions from Ag-NPs,
and this result was comparable to the findings by Kaegi et al. (2011). However, when
studied in activated sludge, it was proposed that the Ag-NPs apparently delivered Ag+ to
the bacteria more effectively, and result in pronounced microbial population shifts that
would hinder some wastewater treatment processes (Yang et al., 2014). ZnO-NPs are
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increasingly and commonly used in cosmetic products, sunscreen formulations, paints,
plastics, and packaging, and are indirectly released into the environment (Ju-Nam and
Lead, 2008). During the anaerobic digestion of cattle manure, the inhibition of methane
production by ZnO-NPs can be partially attributed to the soluble bioavailable fraction of
the metal found in the liquid phase of the reaction after a 14-day incubation period,
although the high toxicity of ZnO-NPs cannot only be explained by the release of toxic
Zn2+
ions (Luna-delRisco et al., 2011). Investigating the chemical transformation of two
ZnO-NPs and one hydrophobic ZnO-NP commercial formulation (used in personal care
products) in anaerobic digestion, Lombi et al. (2012) reported that both “native” Zn, and
Zn added either as a soluble salt or as NPs were rapidly converted to sulfides in all
treatments, which suggests that released Zn2+
ions cannot be the key mechanism for
inhibition of anaerobic digestion by ZnO-NPs. However, Mu and Chen (2011) proposed
that the toxic effect of ZnO-NPs on methane production was mainly due to the release of
Zn2+
from ZnO-NPs, which may inhibit the hydrolysis and methanation steps of digestion.
The negative influence of a shock load of ZnO-NPs on methane production in anaerobic
granular sludge (AGS) has also been reported (Mu et al., 2012). In the same study, a
decrease of proteins in the extracellular polymeric substances (EPS) released by AGS by
69.6% was observed when the dose of ZnO-NPs was greater than 100 mg/g-TSS,
suggesting that the decline of EPS induced by ZnO-NPs resulted in their deteriorating
protective role on the inner microorganisms of AGS, which corresponded with the lower
general physiological activity of AGS observed, and the death of microorganisms.
Besides nanoparticles, carbon nanotubes (NTs) are being investigated for medical
applications because of their theoretical capability to penetrate cell membranes, although
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their low solubility makes them challenging to work with in biological systems. Polar
functional groups, such as polyethylene glycol chains (PEG), are introduced into NTs to
increase their solubility and their bioavailability. Examining the effect of C60 fullerene
NTs on the structure and function of the microbial community in digester sludge, Nyberg
et al. (2008) proposed that C60 fullerenes NTs have no significant effect on the anaerobic
community over a few months exposure according to the community structure changes
monitored by denaturing gradient gel electrophoresis (DGGE), using primer sets
targeting the small subunit rRNA genes of bacteria, archaea, and eukarya. However, in an
anaerobic environment it was found that single-walled NTs (SWNT) functionalized with
polyethylene glycol chains (SWNT- PEG) showed a significant effect on microbial
community structure and function after exposure over a few months by monitoring
methanogenesis, functional gene primers for mcrA gene, and PEG diol dehydratase assay
(Nyberg et al., 2009). When studying the effect of double walled carbon NTs (DWCNT)
on a trichloroethylene (TCE)-dechlorinating culture, it was reported that the rate of
dechlorination first increased as the DWCNT loading was increased, but then decreased
as the DWCNT loading increased beyond 800 mg/L; meanwhile, it was observed that
methanogenesis decreased with increasing DWCNT loading, and no methane was
produced in cultures with 1600 mg DWCNT/L.
2.3.2 Mechanisms of toxicity/inhibition of nanoparticle/nanotubes
As nanomaterials often exhibit physical and chemical properties significantly different
from those of their molecular or macrosized analogs, concern has been growing regarding
their toxicity in the environment (both natural and manmade), and the detailed
mechanisms of how they inhibit anaerobic digestion is still being investigated. One
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important mechanism is the release of metal ions caused by corrosion and dissolution of
the NPs. By investigating algae, crustaceans, fish, bacteria, yeast, nematodes, protozoa
and mammalian cell lines, several studies proposed that the toxicity of CuO NPs is due to
the release of soluble Cu ions (Bondarenko et al., 2013). Apart from the release of toxic
Cu ions, CuO NPs have also been shown to cause toxicity by membrane disruption in
Escherichia coli as well (Zhao et al., 2013). Meanwhile, the toxicity of CuO NPs to
acetoclastic methanogenic activity is most likely caused by both the CuO NPs themselves
and copper ions released to the culture medium (Otero-González et al., 2014). However,
Gonzalez-Estrella et al. (2013) suggested that release of toxic metal species by NP-
dissolution was the principal mechanism of methanogenic inhibition caused by Cu0, CuO,
and ZnO NPs with IC50 values of 62–250 mg/L. Another main mechanism for NPs
toxicity is the generation of reactive oxygen species (ROS) which primarily damage cell
membranes; however, it is unlikely that ROS will be the cause of membrane damage in
anaerobic environments. It is suggested that Al2O3 NPs induced changes in the bacterial
membranes of the anaerobe Ruminococcus flavefaciens 007C by direct physical
interaction, however, under the same conditions nano-TiO2 did not show a significant
effect on the membrane profile of the same bacterium (Vodovnik et al., 2012). On the
other hand, although toxicity of many types of nanotubes to anaerobic microorganisms
have been reported, e.g. DWCNT’s effect on methanogenesis (Kannepalli et al., 2008),
the mechanism of their toxicity is still unknown. A study of the effective visible-light-
driven bismuth vanadate NT for inactivation of bacteria, especially under anaerobic
conditions, indicated that the destruction process of bacterial cells began from the cell
wall to other cellular components, possibly by the hydroxyl radical adsorbing onto the
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surface rather than by the free radical in bulk solution, although both were derived from
photogenerated h+ (Wang et al., 2012). Systematic exposition of NT and NP toxicity to
anaerobic digestion so far is not available. Physical and chemical transformations of
nanomaterials control their toxicity and bioavailability, and therefore must be considered
in future risk assessments.
3 Conclusions
As an efficient waste treatment technology that harnesses natural anaerobic
decomposition to treat waste, reduce waste volume and generate biogas as well,
anaerobic digestion has been widely used as a source of renewable energy. However,
anaerobic digestion can be inhibited to varying degrees by toxic materials present in the
system; these substances may be components of the influent waste stream, or byproducts
of the metabolic activities of the digester bacteria. Inhibitory toxic compounds include
organics, ammonia, sulfide, heavy metals, and the emerging nanomaterials, and are often
present in the processing of wastes from agricultural and industrial operations such as
molasses fermentation, petroleum refining and the tanning industries. These toxic
compounds principally obstruct the activities of the sensitive obligate hydrogen
producing acetogens and methanogenic portions of the digester population, as well as
cause retarded methane formation, a decrease in the methane content of biogas, or can
even cause complete failure of methanogenesis. However, because of the difference in
anaerobic microorganisms and waste composition, results from previous studies on
inhibition of anaerobic processes vary substantially. In this review, we have summarized
and highlighted the effect of specific organic toxicants (e.g., chlorophenols, halogenated
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aliphatics and long chain fatty acids), inorganic toxicants (e.g., ammonia, sulfide and
heavy metals) and the emerging nanomaterials on anaerobic digestion. In addition, better
understanding the mechanism(s) of inhibition or toxicity of different toxicants in an
anaerobic digester provides insights into overcoming these toxic effects and possible
solutions or strategies to properly cope with it, successfully apply anaerobic digestion and
significantly improve waste treatment efficiency. On the other hand, measuring the
toxicant concentration and monitoring them are an essential precautionary strategy. At
present for anaerobic waste treatment, no work has been carried out on the pre-
measurement of toxicity before the waste is introduced into the digester. Most research
has focused on detoxification after inhibition and not on stopping/ameliorating toxicity
before it happens. Hence, a rapid response method to determine toxic substances in the
feed stream, and toxic byproducts in the digester, needs to be developed to protect the
anaerobic microcosm from instability, and hence enable digesters to operate without toxic
perturbations. In addition, how to control toxicity once it has occurred is another
important question to be considered; we will review the state of the art on toxicity
monitoring and control for anaerobic digestion in our next article.
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Table 1
A summary of research findings for the impact of ammonia on anaerobic digestion
Substrate Temperature
(°C)
pH Critical TAN conc. or as
specified (mg/L)
Critical FAN conc. or as
specified (mg/L)
Reference
Synthetic
wastewater
(glucose)
35 6.8-7.0 2480 (50% inhibition methane
producing) -
(Zhou and Qiu,
2006)
Cattle manure 55 7.2
4000 (50% inhibition of aceticlastic) 280 (50% inhibition of
aceticlastic)
(Borja et al., 1996) 7500 (50% inhibition of
hydrogenotrophic methanogens)
520 (50% inhibition of
hydrogenotrophic methanogens)
Synthetic
wastewater
(glucose)
52 7.8 7000 (75% inhibition methane
producing)
620 (100% inhibition methane
producing)
(Siles et al., 2010)
Synthetic
wastewater (yeast
extract)
35 7.7 1445 (50% inhibition methane
producing)
27 (50% inhibition methane
producing)
(He et al., 2011)
Synthetic
wastewater (yeast
extract)
35 8.1 - 800 (COD removal efficiencies
of 78–96%)
(Calli et al., 2005)
Slaughterhouse
by-products 55 7.5
5600 (50% inhibition methane
producing)
635 (50% inhibition methane
producing)
(Bayr et al., 2012)
Swine waste 35 7.6 > 5200 (100% inhibition methane
producing)
200 (100% inhibition methane
producing)
(Garcia and
Angenent, 2009)
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Table 2
IC50 values of six heavy metals and comparison
Bioactivity
parameter
Temperature
(°C) Product
Carbon
source
IC50 (mg/L)
Reference Cd Cr Cu Ni Zn Pb
Methane
production
potential
35 Methane Glucose 36 27 N.A. 35 7.5 - (Altaş, 2009)
Methane
production
potential
37 Methane Whey - - - - 19.2 27.7 (Zayed and Winter,
2000)
Methane
production
potential
35 Methane VFA 330 250 130 1600 270 8000 (Lin and Chen, 1999)
Sulfate-
Reducing
Bacteria
inhibition
36 Sulfide Yeast extract - - 1136 - 1648 - (Utgikar et al., 2003)
Hydrogen
production
potential
26 Hydrogen Sucrose 3300 3000 30 1300 1500 >5000 (Li and Fang, 2007)
Hydrogen
production
potential
35 Hydrogen Dairy
wastewater 170 72 65 - 120 -
(Yu and Fang, 2001a)
(Yu and Fang, 2001b)
Hydrogen
production
potential
37 Hydrogen Glucose - - 350 - >500 - (Zheng and Yu, 2004)
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Fig. 1. Schematic representation of the main conversion processes in anaerobic digestion
(Amaya et al., 2013) including: 1) hydrolysis, breakdown of complex organics into
soluble monomers; 2) acidogenesis, converting small organic molecules into volatile fatty
acids; 3) acetogenesis, converting volatile fatty acids into acetic acid, carbon dioxide, and
hydrogen; and 4) methanogenesis, consuming hydrogen and converting acetate into
methane and carbon dioxide.
Fig. 2. Toxicity sequence of various chlorophenols towards the syntrophic methanogenic
reaction. The logarithm of the partition coefficient of each chlorophenol in octanol/water
(logP) has been listed beside it according to www.guidechem.com.
Fig. 3. A working model showing the direct and indirect inhibition of methanogenesis by
chloroform, trichloroethylene, and perchloroethylene. Presumably, there is an equilibrium
between the protein-bound corrinoids/porphinoids (pool B) and the free
corrinoids/porphinoids (pool F) within methanogen cells to explain the direct and indirect
inhibition of methanogenesis (Yu and Smith, 2000).
Fig. 4. Inhibitory effect of LCFA adsorption over anaerobic granular biomass and
prevention by synthetic adsorbent (bentonite) addition.
Fig. 1.
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Fig. 2.
2.15 2.39 2.50
2.75 2.84 3.06 3.06 3.33 3.62
3.61 3.69 3.72 3.77 4.01 4.56 5.12
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Fig. 3.
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Fig. 4.
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Highlights
critical review of anaerobic toxicity focussing on fundamental mechanisms
looks at specific organics -chlorophenols, halogenated aliphatics, long fatty acids
looks at specific inorganics -ammonia, sulfide and heavy metals
looks at novel nanomaterials and how they inhibit anaerobes
need to develop toxicity sensors with rapid response times