flooded paddy soils (Xu et al. 2015). Similarly, You et al.
(2017) studied the effect of inorganic ENPs on soil enzyme
activities (such as phosphatase and urease) and microbial
communities of alkaline soils. The study observed a considerable change in abovementioned properties along with
harmful impact on biological nitrogen fixation. In another
study, Fe 3 O 4 ENPs at higher concentration significantly
decreased the bacterial count in soil (Jiling et al. 2016).
Similarly, zinc oxide and CeO 2 ENPs considerably affected
various bacterial groups (such as azotobacter, phosphorous,
and potassium solubilizing bacteria) and inhibited various
enzymatic activities (Chai et al. 2015). TiO 2 has shown to
rapidly decline the soil active bacteria and enzymatic
activity, affecting soil microbial characteristics such as
activity, abundance, and diversity (Buzea et al. 2007). In a
similar study, Concha-Guerrero et al. (2014) observed that
CuO ENPs unleashed similar, but a relatively more toxic
impact on soil microbial community. It has been generally
observed that ENPs of inorganic nature have a relatively
greater toxicity than organic ENPs on soil microbial characteristics (Frenk et al. 2013).
In a functional study, CuO and Ag ENPs have shown
reduction in decomposition of leaf (Pradhan et al. 2011). Ag
ENPs, used in a variety of consumer products due to its
antimicrobial properties, significantly impact soil microbial
functional and genomic diversity (Samarajeewa et al. 2017).
However, contrasting studies also exists in the literature (de
Oca-Vásquez et al. 2020). The soil enzymatic activities have
also shown a drastic reduction at high concentrations of
ENPs (Josko et al. 2014; Asadishad et al. 2018). The impact
of ENPs show significant variation with type and dose of
NPs as well as soil properties (Xin et al. 2020). Moreover,
these ENPs have shown negative impact on self-cleaning
ability as well as nutrient providing capacity of soil systems,
which determines the level of plant nutrition and soil fertility
(Suresh et al. 2013). In a manner, soil properties also
determine the toxicity of ENPs. For example, soil pH, textural composition, structure/aggregation, and organic content
affect the soil microbial community and the capability of
these ENPs to unleash toxic effects on soil microorganisms
(Fierer and Jackson 2006; Simonin and Richaume 2015;
Read et al 2016). On the contrary, nanoparticles have also
been termed as “remediation of the future” owing to their
significant role in soil remediation (Sarkar et al. 2019).
5 Nanoparticle’s Toxicity on Environment
The invisible pollution due to ENPs is considered as the
most complicated type of pollutant to control, owing to its
size. The ever-increasing applications and concentrations of
ENPs pose enormous threat of their release into the environment, whose risk assessments are very difficult to
quantify and understand at present stage (Servin and White
2016). The existing literature on eco-toxicological impact of
nanoparticles is somewhat contradictory; however, in general, low to moderate toxicity of these nanoparticles on
terrestrial plants has been observed in most of the scientific
studies. A large number of research studies have focused on
the toxicity assessment of the ENPs used in industries (Du
et al. 2017; Tripathi et al. 2017a, b, c). Generally, the effect
of ENPs on crops (such as spinach, onion, coriander, rice,
wheat, soybean, lettuce, radish, barley, cucumber) has
shown considerable inhibition of seed germination, reduction in shoot and root growth, toxicological effects,
decreased photosynthesis and chlorophyll concentration
(Tripathi et al. 2017a; b, c, d, e). The toxicity level of a
nanoparticle primarily depends upon its solubility and
specificity in binding to the biological site. ENPs of metallic
nature are primarily antimicrobial in nature (Aziz et al. 2016;
Patra and Baek 2017) and show toxicity on the plant cells,
depending on surface charge at the membrane (electrostatic
interaction), which follows the order: mold > yeast > Gramnegative > Gram-positive. Thus, it may unleash an entirely
unknown cascade of change in microbial community
dynamics in the concerned ecosystems, which may turn
lethal on humans in return (Fig. 2).
Carbon-based nano-materials (nanotubes and fullerenes)
can be degraded easily under a wide range of conditions;
however, fullerene is preferably absorbed by wood decaying
fungi and metabolized. As an effect, fullerene nanoparticles
accumulate in microbial cells and are transferred across the
food chain further, owing to feeding relationships (Warheit
et al. 2004). In case of no acute toxicity, bioaccumulation
and long-term exposure to these ENPs may have unforeseen
effects on food chain/web. Similarly, the uptake, accumulation, and build-up of nanoparticles vary in plants,
depending on its type and size, as well as the plant composition. Among the metal-based NMs studied in this regard
(e.g., TiO 2 , Fe 3 O 4 , CeO 2 , ZnO, Ag, Au, Fe, and Cu), only
fullerene and fullerols show a ready uptake tendency in
plants. These NMs enter plant cells variously via aquaporins
(a carrier protein), ion channels, endocytosis, and formations
of entirely new pores across the plant cells, following
apoplastic and symplastic movement and via xylem and
phloem. Remarkably, seed, flower, and fruit strongly import
fluid from the phloem (i.e., sink activity) and have greater
tendency to accumulate ENPs, in relatively higher concentration. Besides toxicological impact on the plant, it raises
issue of safety in human and animal consumption of such
plant organs (Pérez-de-Luque 2017). In all these cases, they
might enter the food chain to unleash unforeseen consequences. Similarly, the excess Fe 3 O 4 nanoparticles produce
some oxidative stress in plant system, affecting photosynthesis, leading to decline in metabolic process rates. ZnO
10
P. Srivastava et al.
(2017) studied the effect of inorganic ENPs on soil enzyme
activities (such as phosphatase and urease) and microbial
communities of alkaline soils. The study observed a considerable change in abovementioned properties along with
harmful impact on biological nitrogen fixation. In another
study, Fe 3 O 4 ENPs at higher concentration significantly
decreased the bacterial count in soil (Jiling et al. 2016).
Similarly, zinc oxide and CeO 2 ENPs considerably affected
various bacterial groups (such as azotobacter, phosphorous,
and potassium solubilizing bacteria) and inhibited various
enzymatic activities (Chai et al. 2015). TiO 2 has shown to
rapidly decline the soil active bacteria and enzymatic
activity, affecting soil microbial characteristics such as
activity, abundance, and diversity (Buzea et al. 2007). In a
similar study, Concha-Guerrero et al. (2014) observed that
CuO ENPs unleashed similar, but a relatively more toxic
impact on soil microbial community. It has been generally
observed that ENPs of inorganic nature have a relatively
greater toxicity than organic ENPs on soil microbial characteristics (Frenk et al. 2013).
In a functional study, CuO and Ag ENPs have shown
reduction in decomposition of leaf (Pradhan et al. 2011). Ag
ENPs, used in a variety of consumer products due to its
antimicrobial properties, significantly impact soil microbial
functional and genomic diversity (Samarajeewa et al. 2017).
However, contrasting studies also exists in the literature (de
Oca-Vásquez et al. 2020). The soil enzymatic activities have
also shown a drastic reduction at high concentrations of
ENPs (Josko et al. 2014; Asadishad et al. 2018). The impact
of ENPs show significant variation with type and dose of
NPs as well as soil properties (Xin et al. 2020). Moreover,
these ENPs have shown negative impact on self-cleaning
ability as well as nutrient providing capacity of soil systems,
which determines the level of plant nutrition and soil fertility
(Suresh et al. 2013). In a manner, soil properties also
determine the toxicity of ENPs. For example, soil pH, textural composition, structure/aggregation, and organic content
affect the soil microbial community and the capability of
these ENPs to unleash toxic effects on soil microorganisms
(Fierer and Jackson 2006; Simonin and Richaume 2015;
Read et al 2016). On the contrary, nanoparticles have also
been termed as “remediation of the future” owing to their
significant role in soil remediation (Sarkar et al. 2019).
5 Nanoparticle’s Toxicity on Environment
The invisible pollution due to ENPs is considered as the
most complicated type of pollutant to control, owing to its
size. The ever-increasing applications and concentrations of
ENPs pose enormous threat of their release into the environment, whose risk assessments are very difficult to
quantify and understand at present stage (Servin and White
2016). The existing literature on eco-toxicological impact of
nanoparticles is somewhat contradictory; however, in general, low to moderate toxicity of these nanoparticles on
terrestrial plants has been observed in most of the scientific
studies. A large number of research studies have focused on
the toxicity assessment of the ENPs used in industries (Du
et al. 2017; Tripathi et al. 2017a, b, c). Generally, the effect
of ENPs on crops (such as spinach, onion, coriander, rice,
wheat, soybean, lettuce, radish, barley, cucumber) has
shown considerable inhibition of seed germination, reduction in shoot and root growth, toxicological effects,
decreased photosynthesis and chlorophyll concentration
(Tripathi et al. 2017a; b, c, d, e). The toxicity level of a
nanoparticle primarily depends upon its solubility and
specificity in binding to the biological site. ENPs of metallic
nature are primarily antimicrobial in nature (Aziz et al. 2016;
Patra and Baek 2017) and show toxicity on the plant cells,
depending on surface charge at the membrane (electrostatic
interaction), which follows the order: mold > yeast > Gramnegative > Gram-positive. Thus, it may unleash an entirely
unknown cascade of change in microbial community
dynamics in the concerned ecosystems, which may turn
lethal on humans in return (Fig. 2).
Carbon-based nano-materials (nanotubes and fullerenes)
can be degraded easily under a wide range of conditions;
however, fullerene is preferably absorbed by wood decaying
fungi and metabolized. As an effect, fullerene nanoparticles
accumulate in microbial cells and are transferred across the
food chain further, owing to feeding relationships (Warheit
et al. 2004). In case of no acute toxicity, bioaccumulation
and long-term exposure to these ENPs may have unforeseen
effects on food chain/web. Similarly, the uptake, accumulation, and build-up of nanoparticles vary in plants,
depending on its type and size, as well as the plant composition. Among the metal-based NMs studied in this regard
(e.g., TiO 2 , Fe 3 O 4 , CeO 2 , ZnO, Ag, Au, Fe, and Cu), only
fullerene and fullerols show a ready uptake tendency in
plants. These NMs enter plant cells variously via aquaporins
(a carrier protein), ion channels, endocytosis, and formations
of entirely new pores across the plant cells, following
apoplastic and symplastic movement and via xylem and
phloem. Remarkably, seed, flower, and fruit strongly import
fluid from the phloem (i.e., sink activity) and have greater
tendency to accumulate ENPs, in relatively higher concentration. Besides toxicological impact on the plant, it raises
issue of safety in human and animal consumption of such
plant organs (Pérez-de-Luque 2017). In all these cases, they
might enter the food chain to unleash unforeseen consequences. Similarly, the excess Fe 3 O 4 nanoparticles produce
some oxidative stress in plant system, affecting photosynthesis, leading to decline in metabolic process rates. ZnO
10
P. Srivastava et al.