3.4. Effects of ENPs on nitrification during anaerobic ammonium
oxidation
Anaerobic ammonium oxidation (anammox) process is a novel biological
nitrogen removal technology that is gaining popularity for nitrogen
removal in wastewater streams. In this process, ammonium is directly
converted to dinitrogen gas using nitrite as the electron acceptor in
the absence of oxygen (Eq. 5) for nitrogen removal from wastewater
streams.
\begin{equation}
\text{NH}_{4}^{+}+1.32\text{NO}_{2}^{-}+0.066\text{HCO}_{3}^{-}+0.13H^{+}\rightarrow 1.02N_{2}+0.26\text{NO}_{3}^{-}+0.066\text{CH}_{2}O_{5}N_{0.15}+2.03H_{2}O\ \ \ \ (Eq.5)\nonumber \\
\end{equation}
No addition of organic carbon source is required since
CO2 is utilized as the only carbon source. In addition,
it significantly reduces oxygen demand since ammonium is only required
to be partially nitrified to NO2-instead of NO3-; thus leading to
considerable saving in operational cost. Due to these and other
advantages (low CO2 emission and low biomass yield),
anammox processes have been widely regarded as an innovative and
sustainable alternative to the classical activated sludge process (Zhang
ZZ et al., 2016). There are more than one hundred full-scale anammox
installations worldwide that are being applied for the treatment of
side-stream wastewater (reject water).
A few studies have studied addressing the of ENPs effects on nitrogen
removal by annamox processes (Table S4). Zhang et al (2017 ) studied the
short-term (24 hrs) effects of CuNPs, CuONPs, ZnONPs and AgNPs in a
batch study using anammox sludge. Their results showed that CuONPs,
ZnONPs and AgNPs up to 50 mg/g suspended solid did not affect anammox
activity, ROS generation or LDH release. By contrast, CuNPs at 1.25 and
2.5 mg/g SS resulted in severe inhibition of anammox activity, without
inducing an increase in LDH release. Higher loads of CuNPs caused
significant inhibition of anammox activity and increased LDH release.
The toxicity was primarily attributed to dissolved
Cu2+ ions. Another batch study (Zhang et al. 2017)
demonstrated that the addition of EDTA or S2- could
attenuate the adverse effects of CuNPs, presumably due to the chelation
or sulfidation of Cu2+ ions. Later, the long-term
effect of CuNPs was studied by adding CuNPs to an up-flow anaerobic
sludge blanket (UFAB) reactor at 0.5 mg/L for 15 days, 1.0 mg/L for 15
days, and 5 mg/L for 30 days. Results showed that 5 mg/L of CuNP caused
near complete inhibition of nitrogen removal and significant a decrease
of the abundance of anammox bacteria. Withdrawing CuNPs from the
influent permitted the recovery of nitrogen removal.
The long-term effects of ZnO NPs on annammox sludge was also studied
using UFAB reactor (Zhang et al. 2018). ZnO NPs (~30 nm)
were added to the bioreactor at 1.0 mg/L on day 31, increased to 5.0
mg/L on day 46 and 10 mg/L on day 61. Results showed that shock-load of
10 mg/L ZnONPs resulted in the deprivation of 90% of the nitrogen
removal capacity within 3 days. Anammox activity was significantly
inhibited without any significant increase in LDH release or
intracellular ROS production. These effects were attributed to dissolved
Zn2+ ions and complete recovery was observed within 40
days after withdrawing the NPs from the influent. Another study by the
same group investigated the effects of other metal oxides NPs on the
performance of anammox process (Zhang et al. 2018). SiO2 NPs
(~30 nm), TiO2 NPs (~60 nm,
hydrophilic), CeO2 NPs and
Al2O3 NPs (30 nm, hydrophilic) on
granular anammox sludge in lab-scale UFAB bioreactors. NPs were added to
the bioreactors at 1, 50 and 200 mg/L in a step-wise fashion with a
30-day interval and lasted for an entire duration of 90 days. No adverse
effects on nitrogen removal were observed, and this resilience was
attributed to adaptation of the microorganisms through community shift
and enhanced EPS production. Most recently, Li et al (Li et al. 2019 )
reported that exposure to graphene oxide (1 and 10 mg/L) resulted in
acute toxicity and inhibition of annamox nitrogen removal. The effects
disappeared by day 19 and reversed by the end of the study at day 61,
with a TN removal efficiency higher than control. The same doses of AgNP
caused long-term inhibition on TN removal, which did not recover. The
long-term enhancement of TN removal by GO was accompanied by the
relative high abundance of anammox bacteria C.Anammoxoglobus ;
while the TN removal inhibition by AgNP was accompanied by the
disappearance of some species with anammox ability. This observation
seems to contradict with the findings by Zhang et al (2018), in which
they reported no long-term adverse effects of AgNPs on anammox activity
at concentrations of 1, 10 and 50 mg/L. This discrepancy may be related
to differences in the type of sludge, bioreactor and particles used in
these studies.
The effects of iron NPs seemed to be beneficial to anammox nitrogen
removal. Li et al (2018) reported that adding Fe3O4 NPs (1, 10 mg/L) to
an unplanted anammox subsurface flow constructed wetlands produced
concentration-dependent acute toxic effects on ammonia removal; these
effects disappeared overtime and by day 61, nitrogen removal rate were
actually enhanced. Nano scale zero valent iron (nZVI) have also been
proved to be beneficial for anammox bacteria growth and nitrogen removal
(Erdim et al. 2019). In summary, these early studies have shown that in
an anammox process, ENP toxicity was mainly caused by dissolved ions;
the role of ROS generation was less significant than in the conventional
activated sludge process, likely due to the lack of oxygen supply.
4. Mechanisms of ENP toxicity
Several mechanisms for ENP toxicity have been proposed based on
experimental observations (Figure 1). Metal based ENPs are believed to
exert their toxicity mainly through dissolved ions, in combination with
the effects from nanoparticles. Metal ions bind with the negatively
charged compounds in the bacteria cell wall, resulting in cell wall
destabilization or collapse. Metal ions have high affinity to molecules
containing –SH groups, such as cysteine; this binding can break S-S
bond bridges that are necessary to maintain the integrity of folded
proteins or directly disrupt the function of certain enzymes (Slavin et
al. 2017). For example, the activity of most AMOs in N. Europaeaare inhibited by Co2+, Zn2+,
Ni2+, and Fe2+ in a
concentration-dependent manner (Ensign et al. 1993). The dissolution of
metal-based NPs also generates reactive oxygen species, which could
cause cell membrane damage via oxidation of membrane components such as
lipids. The internalized metal ion can also react with mitochondrial
H2O2 and produce intracellular ROS or
affect DNA repair and cause mutation (Huangfu et al., 2019).
Intracellularly-produced ROS may also damage the cell membrane via lipid
oxidation or damage cellular DNA without visible cell membrane damage.
Non-metal based ENPs such as carbon nanomaterials also produce ROS upon
light illumination, a property that is shared by TiO2 and ZnO NPs.
ENPs may enter cells through direct penetration or endocytosis. Direct
penetration is caused by non-specific binding forces (electrostatic,
hydrophobic, van der Waals) between the particle and the cell membrane;
while endocytosis involves specific receptor-ligand interactions. Once
inside the cell, ENPs can bind with intracellular biomolecules, interact
with mitochondria, induce ROS production, or damage cellular functions
(Yang et al, 2019). Increased ROS production leads to enzyme
inactivation and DNA damage, likely the reason for the observed
inhibition of key enzymes involved with nitrification.
ENPs can cause physical damage to bacteria in various ways. Adsorption
of ENPs onto cell wall/membrane leads to depolarization of the cell
wall/membrane, which changes the negativity of the membrane and makes it
more permeable. Carbon nanotubes can puncture the cell membrane like
needles. Graphene nanosheets can both cut through cell membrane and also
disruptively extract phospholipids from the membrane due to strong van
der Waals interactions and hydrophobic effects. These types of physical
damage disrupt and weaken the cell membrane, resulting in release of
biomolecules such as LDH. NP aggregation onto the cell surface may also
facilitate NP dissolution, releasing metal ions that can easily enter
the cell (Slavin et al, 2017). Large graphene nanosheets and aggregates
of smaller NPs can entrap bacteria and prevent them from taking up
nutrients. Long CNTs can wrap around the bacteria and induce osmotic
cell lysis.
EPS generally acts as a protective layer for the microorganisms by
absorbing ENPs and the dissolved ions. On the other hand, the EPS may
promote ENP dissolution after the absorption capacity has been reached.
Under some circumstances, strong interactions between ENPs and the EPS
may result in stripping of the protective EPS layer off sludge
microorganisms, thus making the microorganisms more vulnerable. As
mentioned earlier, the ENP-microorganism interactions also depend
heavily on the properties of the latter including type (gram-negative
vs. gram-positive) (Mocan et al, 2017), shape (rod-shaped vs. spherical)
(Al-Jumaili et al, 2017), hardness (Liu et al, 2009), cell wall
structure (lipopolysaccharides, phospholipids defects) (Hsu et al,
2016), and enzyme and metabolism activities (Krishnamoorthy et al.
2012). Therefore, it is expected that different microorganisms in the
activated sludge will respond differently to the same ENP stress,
resulting in microbial community shifts as observed by several studies
cited in this review. These shifts however, may not always cause
inhibition.
5. Issues and challenges
There are a number of issues and challenges associated with assessing
the effects engineered nanomaterials on nitrification in activated
sludge.
First, even though NP physical-chemical properties are critical factors
causing microbial toxicity, current literature is frequently missing
critical data regarding NP physical-chemical properties, including
hydrodynamic size, shape, surface charge, hydrophobicity, surface
roughness, deformability (soft vs hard NPs), surface chemistry,
electronic structure and coating. The most frequently provided data is
particle size, many of them determined with TEM, but the hydrodynamic
size is more appropriate for NPs in an aqueous environment. Some papers
reported the size information provided by the manufacturer, which is not
always correct and requires verification. A limited number of papers
provided zeta potential measurements, even fewer have included
hydrophobicity, surface roughness, deformability (soft vs hard NPs),
surface chemistry, or coating (Tables S1-S3). Lack of comprehensive
physical-chemical characterization makes it very difficult to draw
meaningful comparisons between studies, because NP toxicity and
interactions with the microorganisms and the EPS are governed by these
properties (Huangfu et al., 2019, Slavin et al., 2017).
Second, it is also necessary to understand fate and transformation of
these particles post entry into the wastewater. Majority of these
particles, especially those which are not stabilized with coatings, will
undergo changes and take on a new physio-chemical identity. These
changes could include biodegradation, dissolution, precipitation,
aggregation, adsorption of naturally occurring substances and chemical
transformation, depending on both particle-specific properties &
particle state (free or matrix incorporated), and on the chemistry of
the surrounding solution (pH, ionic strength, ionic composition) (Petosa
et al., 2010). As a result, NPs identity in the wastewater could be
vastly different from the original NPs. Clar et al (Clar et al., 2016)
showed that aggregated CuO NPs in wastewater were about 1600 nm in
diameter, about 40 times larger than the original NP (46 nm). Since it
is these transformed particles that interact directly with the
microorganisms, a thorough characterization of these transformed NPs is
vital.
NP transformations are particularly relevant in the pipes that carry
sewage to wastewater treatment plants. This underground network is
anaerobic and contains soluble and insoluble constituents that may react
with a wide range of ENPs (Metcalf and Eddy, 2014). The principle
compound of interest is sulfide, which is biologically produced by a
consortia of sulfate-reducing microorganisms. Sulfide forms complexes
with Ag NPs (Kaegi et al., 2013), Cu NPs (Hatamie et al., 2014), and ZnO
NPs (Lupitskyy et al., 2018). There are also coarse particles and
colloids that can heteroaggregate with ENPs during transit in the sewer
(Zhang et al., 2016). These processes tend to reduce the bioavailability
and ecotoxicity of ENPs, however, the extent of NP transformation may be
limited in sewer systems with short residence times.
There are other important, unresolved issues. The synergetic effects of
multiple species of NPs are particularly relevant for WWTPs due to the
presence of a vast variety of NPs found in sewage. The purity of the
nanomaterials and variations among different preparations must also be
considered; this is important because the toxicity may be affected by
the impurities during the manufacturing process. For instance, CuO NPs
that contains Cu NPs contaminants are expected to have a toxicity
profile different from that of pure CuO NPs, since the release of
dissolved Cu2+ions from CuO NPs is significantly
slower than from CuNPs (Zhang et al., 2017 ). The majority of literature
so far has used unfunctionalized or minimally functionalized ENPs in
their studies, while in reality, many applications use functionalized
ENPs because of the improved efficacy, usability or added functionality.
Functionalized ENPs will have different surface structure, chemistry,
and aggregation properties; all of which will result in a completely
different toxicity profiles.
Lastly, there is considerable uncertainty about the values of ENP
concentrations that are present in wastewater. There are a small number
of published studies that present such data and the range and temporal
variability of ENP concentrations in domestic wastewater are not yet
understood. There is a major information gap because it is difficult to
relate published findings to realistic operational scenarios. There is a
need to determine environmentally relevant concentrations of ENPs using
field sampling campaigns. Such work should be done with a well-organized
series of grab samples, taken together with flow and water quality data.
6. Conclusions
There is an urgent need to understand the effects of ENPs on wastewater
treatment because of the growing use of nanomaterials. Nitrification is
a critical wastewater treatment process that may be disrupted under
certain conditions. Studies have confirmed that short-term,
environmentally-relevant concentrations generally did not inhibit
nitrification in conventional activated sludge systems. Long-term
exposure to relatively high concentrations of ENPs inhibited nitrogen
removal and shifted the microbiological community structure in activated
sludge. Some studies have shown resiliency of activated sludge systems.
Physical & chemical properties of the ENP, properties of the
microorganisms and their environment are all believed to contribute to
the variabilities in toxicity results observed in literature. Several
mechanisms may contribute to the ENP-induced toxicity, including
physical disruption of the cell membrane, generation of ROS, inhibition
of enzymes and metabolic processes, and intracellular accumulation of
ENPs. The effects of non-metal and composites ENPs have not been
well-studied and need to be thoroughly investigated in future studies;
as these materials are gaining increasing popularity in real
applications. Aggregation and transformation of ENPs in wastewater are
common and thus the observed toxic effects are in fact caused by
aggregates or transformed NPs, thorough characterization of these
“transformed NPs” will help to better interpret the results and
explain the variabilities among different studies. Early studies on the
emerging anammox technology provided evidences of ENP-induced
nitrification inhibition in these processes, however, the mechanisms are
expected to different from that in activated sludge due to the lack of
oxygen and differences in the nitrification microorganisms. Future
research should also include the even more recent development of
biological nitrogen removal processes that combine partial nitrification
with anammox.