Removal of Cr(VI) from water by in natura and magnetic nanomodified hydroponic lettuce roots

Abstract

Biosorption is a viable and environmentally friendly process to remove pollutants and species of commercial interest. Biological materials are employed as adsorbents for the retention, removal, or recovery of potentially toxic metals from aqueous matrices. Hexavalent chromium is a potential contaminant commonly used in galvanoplasty and exhibits concerning effects on humans and the environment. The present work used in natura lettuce root (LR) and nanomodified lettuce root (LR-NP) for Cr(VI) adsorption from water medium. The nanomodification was performed by coprecipitation of magnetite nanoparticles on LR. All materials were morphologically and chemically characterized. The conditions used in removing Cr(VI) were determined by evaluating the pH at the point of zero charge (pHPZC = 5.96 and 6.50 for LR and LR-NP, respectively), pH, kinetics, and sorption capacity in batch procedures. The maximum sorption capacity of these materials was reached at pH 1.0 and 30 min of adsorbent-adsorbate contact time. The pseudo-second-order kinetic equation provided the best adjustments with r² 0.9982 and 0.9812 for LR and LR-NP, respectively. Experimental sorption capacity (Qexp) results were 4.51 ± 0.04 mg/g, 2.48 ± 0.57 mg/g, and 3.84 ± 0.08 mg/g for LR, NP, and LR-NP, respectively, at a 10 g/L adsorbent dose. Six isothermal models (Langmuir, Freundlich, Sips, Temkin, DR, and Hill) fit the experimental data to describe the adsorption process. Freundlich best fit the experimental data suggesting physisorption. Despite showing slightly lower Qexp than LR, LR-NP provides a feasible manner to remove the Cr(VI)-containing biosorbent from the medium after sorption given its magnetic characteristic.

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Environmental Science and Pollution Research
https://doi.org/10.1007/s11356-022-21755-0
EFFECTIVE WASTE MANAGEMENT WITHEMPHASIS ONCIRCULAR ECONOMY
Removal ofCr(VI) fromwater byin natura andmagnetic nanomodified
hydroponic lettuce roots
BeatrizCalimanSoares1· ThaisEduardaAbilio1· JuliaCristinaJosé1· GeórgiaLabuto2,3·
ElmaNeideVasconcelosMartinsCarrilho1,4
Received: 31 October 2021 / Accepted: 26 June 2022
© The Author(s), under exclusive licence to Springer-Verlag GmbH Germany, part of Springer Nature 2022
Abstract
Biosorption is a viable and environmentally friendly process to remove pollutants and species of commercial interest.
Biological materials are employed as adsorbents for the retention, removal, or recovery of potentially toxic metals from
aqueous matrices. Hexavalent chromium is a potential contaminant commonly used in galvanoplasty and exhibits con-
cerning effects on humans and the environment. The present work used in natura lettuce root (LR) and nanomodified
lettuce root (LR-NP) for Cr(VI) adsorption from water medium. The nanomodification was performed by coprecipitation
of magnetite nanoparticles on LR. All materials were morphologically and chemically characterized. The conditions used
in removing Cr(VI) were determined by evaluating the pH at the point of zero charge (pHPZC = 5.96 and 6.50 for LR and
LR-NP, respectively), pH, kinetics, and sorption capacity in batch procedures. The maximum sorption capacity of these
materials was reached at pH 1.0 and 30min of adsorbent-adsorbate contact time. The pseudo-second-order kinetic equation
provided the best adjustments with r2 0.9982 and 0.9812 for LR and LR-NP, respectively. Experimental sorption capacity
(Qexp) results were 4.51 ± 0.04mg/g, 2.48 ± 0.57mg/g, and 3.84 ± 0.08mg/g for LR, NP, and LR-NP, respectively, at a 10
g/L adsorbent dose. Six isothermal models (Langmuir, Freundlich, Sips, Temkin, DR, and Hill) fit the experimental data
to describe the adsorption process. Freundlich best fit the experimental data suggesting physisorption. Despite showing
slightly lower Qexp than LR, LR-NP provides a feasible manner to remove the Cr(VI)-containing biosorbent from the
medium after sorption given its magnetic characteristic.
Keywords Biosorption· Biomass· Water treatment· Magnetite nanoparticles· Hexavalent chromium
Introduction
The growing industrialization is a substantial cause of envi-
ronmental degradation, mainly by disposing contaminated
effluents inappropriately (Dhankhar and Hooda 2011). The
discarding of effluents with no prior treatment results in the
introduction of pollutants such as potentially toxic metals,
dyes, and drug residues, which are harmful to human health
and aquatic fauna, causing mortality of species existing in
these ecosystems (Santos etal. 2011).
Among the most harmful contaminants are toxic metals,
which can be a severe environmental and health risk when
present in concentrations higher than legally permitted. Met-
als are not biodegradable and can be a health risk due to their
toxicity and carcinogenicity. Several anthropogenic sources
introduce chromium, a very rigid toxic metal in the environ-
ment (Rossi etal. 2018; Jobby etal. 2018).
Communicated by Tito Roberto Cadaval Jr.
* Elma Neide Vasconcelos Martins Carrilho
elma.carrilho@gmail.com
1 Laboratory ofPolymeric Materials andBiosorbents,
Universidade Federal de São Carlos, Araras, SP13600-970,
Brazil
2 Departamento de Química, Universidade Federal de São
Paulo, Diadema, SP09913-030, Brazil
3 Laboratory ofIntegrated Sciences, Universidade Federal de
São Paulo, Diadema, SP09913-030, Brazil
4 Departamento de Ciências da Natureza, Universidade
Federal de São Carlos, Matemática e Educação, Araras,
SP13600-970, Brazil

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Chromium occurs naturally in the earth, but different
from other toxic metals; it shows oxidation states such as
Cr3+ and Cr6+. These species are directly related to pH and
redox conditions in the aqueous phase. Hexavalent chro-
mium is its most toxic form, which is not found naturally in
water but ends up being mainly discharged through effluents
from steel industries (Conceição etal. 2014; Kanagaraj and
Elango 2019). A high concentration of hexavalent chro-
mium causes lung and stomach cancer, diabetes, and nasal
cavity, different fromCr(III), which is an essential micro-
nutrient for humans (Kanagaraj and Elango 2019). Due to
the severe consequences of Cr(VI) in water, several treat-
ment methods have been proposed for its removal. Although
conventional wastewater treatment methods (chemical pre-
cipitation, flocculation, ion exchange, and filtration) are still
employed, their application presents various limitations.
They are not economically feasible, require a long time to
be implemented, and are not doable if contaminants are
found in large volumes of water (Santos etal. 2011; Babu
etal. 2019). Among the alternatives for removing pollut-
ants from the environment, biosorption, a physicochemical
process started in the early 1970s, has been proposed and
differs from other methods due to its low cost and high
efficiency (Abilio etal. 2021; Carvalho etal. 2021; Cid
etal. 2020; Gadd 2009). Biomass retains contaminants on
the surface due to the active sites available in its composi-
tion and can uptake cations and anions by physicochemi-
cal interaction occurring between adsorbate and adsorbent.
Biosorption consists of the use of biomasses from biologi-
cal materials, such as hydroponic lettuce root, a widely
available residue with no economic relevance. Previous
studies using in natura and chemically modified hydroponic
lettuce roots demonstrated their potential as adsorbents for
the removal of Cu(II), Fe(II), Zn(II), and Mn(II) (Milani
etal. 2018a, 2018b).
Lettuce is a widely consumed vegetable. In 2005, world
lettuce production was 22 million tons, and it is mainly culti-
vated in temperate and subtropical regions (Mou 2008). This
vegetable grows a deep taproot with horizontal lateral roots
near the soil and water, absorbing water and nutrients (Mou
2008). For this reason, lettuce roots are promising biomass
to be used as biosorbentsdue to the presence of cellulose,
hemicellulose, and proteins (Akhter etal. 2014).
Some techniques can be employed to improve and facili-
tate the biosorption process or increase the sorption capacity
of the material, such as the use of magnetic nanoparticles.
Due to the small sizes of these materials, which range from
1 to 100nm, they have properties that cause repulsive and
attractive interactions with the magnetic field and can be
easily removed from the medium with the use of a magnet
after the sorption process (José etal. 2019; Debs etal. 2019;
Labuto etal. 2018).
Given the incorrect disposal of wastewater containing emerg-
ing contaminants, such as potentially toxic metals, it is neces-
sary to think of ways to remedy these ecosystems, to reduce the
damage to living beings. Chromium(VI) is a toxic metal and
exhibits carcinogenic effects. Its emission comes mainly from
leather tanning, chromate production, paint manufacturing, and
electroplating (ARIF etal. 2020). The use of lettuce roots as
biosorbents for Cr(VI) in its removal is a new approach since this
material has not yet been used. Therefore, this work proposes
this environmentally friendly, promising, and economically via-
ble alternative to treat Cr(VI)-contaminated water. In addition,
nanomodification of the roots’ biomass facilitates the removal
of the adsorbent from the medium after the adsorption process.
In this way, this research contemplates significant aspects of this
biomass potential that has not been previously reported.
The aim of this work is to prepare in natura and nanomod-
ified hydroponic lettuce roots from domestic and agro-indus-
trial residues and use them for the removal of Cr(VI) species
in the aqueous medium, as a promising alternative for water
and effluent remediation.
Material andmethods
Reagents andsamples
All solutions used in this work were prepared using deion-
ized water obtained from a Direct-Q 3 System (Merck Mil-
lipore, Germany). The buffer solution at pH 5.5 was pre-
pared from 100% glacial acetic acid and potassium acetate
dilutions to obtain a 0.005mol/L KCH3COO/CH3COOH
(Synth, São Paulo-SP, Brazil) solution. The Cr(VI) stock
solution was prepared by dissolving K2Cr2O7 (LabSynth,
São Paulo, Brazil) in deionized water. The magnetic mate-
rials were prepared with FeCl3·6H2O, FeSO4·7H2O, HCl,
and NH4OH, and ethanol was used to wash the obtained
nanomaterials (all from LabSynth, São Paulo-SP,Brazil).
Hydrochloric acid, NaOH, and NaCl (Synth, São Paulo-SP,
Brazil) were used in the pH assessment point of zero charge
(pHPZC) determination procedure.
The hydroponic lettuce root biomass used in the present
work was produced in the Federal University of São Carlos
(Araras-SP, Brazil). This material was washed with distilled
deionized water for the total removal of solid residues. After
washing, the material was oven-dried (Te-394/1, Tecnal
Scientific, Piracicaba-SP), ground in a rotor mill, passed
through a 0.12-mm sieve, and stored.
Synthesis of Fe3O4 nanoparticles andthemagnetic
bionanocomposite (LR‑NP)
A coprecipitation method (Panneerselvam etal. 2011;
Cardona etal. 2019) was used to synthesize magnetite

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nanoparticles (NPs) and the magnetic nanomodified let-
tuce root (LR-NP) biomass. The procedure consists of the
dissolution of Fe(II) and Fe(III) salts at a 1:2 molar ratio,
respectively, in a 1.0mol/L HCl medium. Under constant
stirring, 0.7mol/L NH4OH was slowly added to this solu-
tion for about 30min. After this synthesis, the in natura
lettuce root (LR) was added to the synthesized NP at a 5:1
ratio, respectively, according to Abilio etal. (2021), under
constant heating and stirring to prepare the nanocomposite
LR-NP. Reactions 1 and 2 demonstrate the synthesis of NP
and LR-NP, respectively.
Characterization ofLR, NP, andLR‑NP
All materials were characterized by Fourier transform infrared
spectroscopy (FTIR, Bruker, Vertex Model), scanning elec-
tron microscopy (ZEISS LEO 440 with OXFORD detector
7060 model), and X-ray diffraction (XRD, Rigaku MiniFlex
600 Model). FTIR analysis was used to identify functional
groups responsible for toxic metal removal present in the
surface of the biosorbent (Babu etal. 2019). The equipment
operated with 32 scanners ranging from 4000 to 400 cm−1,
using pellets made with the samples and potassium bromide.
Scanning electron microscope (SEM) analysis was car-
ried out to obtain external images of the material’s surface.
It operated with a 10-kV electron beam, 2.82-A current,
and 200-pA probe. The samples were carbon-coated in a
BAL-TEC MED 020 Coating System metallizer (BAL-TEC,
Liechtenstein) and desiccated until analysis.
The X-ray diffraction analysis allows determining the
crystalline structures present in the materials. A powder
X-ray diffractometer equipped with a copper X-ray source
(CuKα tube) was used with λ = 1.5406Å, 40-kV voltage,
and 30-mA current. Data were obtained in the range of 2°
to 90° at a rate of 0.02° per second.
pH assessment: point ofzero charge andsorption
pH
The pHPZC is the pH in which the charge of the biosorb-
ent surface is electrically zero. The material is positively
charged when exposed to a solution at pH below the pHPZC.
If the solution is at a pH greater than pHPZC, the material’s
surface will be negatively charged. It allows to predict the
material surface charge and conduct the adsorption assays
under a more suitable condition for the adsorption process
(Bakatula etal. 2018; Brito etal. 2019). Thus, it is possible
to get the best pH values to perform adsorption.
(Reaction 1)
Fe2+
(aq)
+2Fe
3+
(aq)
+8OH
−
(aq)→Fe3O4(s) +4H2O
(l)
(Reaction 2)
Fe
3
O
4
(s) + LR(s)
→
LR − Fe
3
O
4
(s)
Firstly, 0.1mol/L NaCl solutions with initial pH values rang-
ing from 2.0 to 12.0 were added to LR or LR-NP at a 1.0g/L
dosage. The suspensions were left under constant stirring at
185rpm for 24h, after which the final pH was measured.
After the pHPZC determination, pH assessment was per-
formed with pH values ranging from 1.0 to 6.0 for LR and
1.0 to 6 for LR-NP, using HCl or NaOH 1mol/L to adjust
Cr(VI) solutions pH. LR and LR-NPat a25 g/L biosorb-
ent dose were added to Falcon tubes containing 10mg/L of
Cr(VI) solution at different pH values. The suspensions were
kept under constant stirring for 10min at 185rpm. Using a
Nd magnet, the supernatant was separated for further analy-
sis by flame atomic absorption spectrometry (FAAS, AAna-
lyst 400, PerkinElmer, USA) employing a 1.0 to 6.0mg/L
Cr(VI) calibration curve, and the operating parameters:
nebulization flow (5.0 L min/L), air (10.0 L/min), acetylene
flow (3.3 L/min), and nebulizer flow rate (2.0 L/min), using
a Cr cathode lamp (357.87nm).
All Cr(VI) sorption assessments (pH, kinetics, and sorp-
tion capacity) were performed using batch procedures
employing the scheme depicted in Figure S1 available in
the Supplementary Material.
Kinetic studies
The kinetic study is significant for optimizing the sorp-
tion process, and the procedure is performed by the reac-
tion between adsorbate and adsorbent. The best contact
time is when equilibrium is reached. Several kinetic mod-
els describe the adsorbate/adsorbent interactions and the
adsorption process in the literature. The adjustment is
performed according to the kinetic model thatbest fit the
experimental data(Milani etal. 2018b; Raganati etal. 2019).
Two kinetic models were used in the present work: the
pseudo-first-order where it assumes that the number of
active sites is proportional to the adsorption rate, and the
pseudo-second-order, which proposes that the adsorption
rate is proportional to the square of the number of active
sites present on the adsorbent surface. The mathematical
equation of the pseudo-first-order is represented in Eq.1 (Ho
2006; Raganati etal. 2019).
where k1 is the constant of pseudo-first-order (min−1);
qe is the sorption capacities in equilibrium (mg/g); qt is the
sorption capacity at a given time of equilibrium (mg/g); t is
the sorption time (min).
Furthermore, the linear equation for the pseudo-second-
order is represented in Eq.2 (Ho and Mckay 2000).
(1)
ln(qe−qt)=ln(qe)−k1
⋅
t
(2)
t
qt
=
1
k2qe
2+
t
q
e
.
t

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where k2 is the constant of pseudo-second-order
(g/(mg/min)); qe is the mass of adsorbate per gram
of adsorbent in equilibrium (mg/g); qt is the mass of
adsorbate per gram of adsorbent in equilibrium, at a
given time (mg/g); and t is the sorption time (min).
The kinetics approach was performed using a
10mg/L Cr(VI) solution with the previously optimized
pH = 1.0 and the biosorbents LR or LR-NP at a 10g/L
dosage. The suspensions were kept under constant
stirring at 185rpm using 5, 10, 15, 20, 25, 30, 60,
90, 120, and 150min contact time between adsorbate
and adsorbent. The supernatants were analyzed for Cr
determination by FAAS under the previously described
operational conditions. This procedure was performed
in triplicate.
Sorption capacity andisotherm model parameters
Chromium(VI) solutions were prepared at increasing con-
centrations (10 to 100mg/L) at a previously optimized
pH value. The solutions were mixed with LR, NP, and
LR-NP at a 10g/L adsorbent dosage. The suspensions
were kept under stirring (185rpm) for 30min at the best
sorption pH (1.0). Thus, the supernatants were removed
and analyzed by FAAS to determine the remaining Cr
contents under the same conditions previously described
in the pH assessment section.
In the experimental sorption capacity assessments, the
amount of Cr adsorbed for each concentration (Qe) was
calculated after reaching equilibrium. It was determined
by Eq.3, taking into account Cr concentration in the work
solutions before and after sorption, the adsorbent mass,
and the adsorbate volume (José etal. 2019).
where Qe is the concentration of adsorbed Cr(VI), in
equilibrium (mg/g); C0 is the initial Cr(VI) concentration
(mg/L); Ce is the Cr(VI) concentration in solution, after
sorption (mg/L); V is the volume of Cr(VI) solution (L);
and m is the biosorbent mass (g).
Six isotherm models, Freundlich, Temkin, Sips, Lang-
muir, Hill, and D-R, were applied to the experimental data
of Cr(VI) sorption to evaluate the best possible adjust-
ment. This analysis correlates the mechanism of adsorp-
tion of solutions on solid surfaces with the adsorption
isotherm models (Giles etal. 1960). The equations of the
non-linear isotherm models used in the current study to
assess Cr(VI) sorption by magnetic NPs, LR, and LR-NP,
and their respective parameters, are presented in TableS1
(Supplementary Material).
(3)
Q
e=
(
C0−Ce
)
V
m
Results anddiscussion
Characterization ofNP, LR, andLR‑NP
The FTIR technique was used to identify the main func-
tional groups presented inLR and LR-NP before and after
chromium sorption (LR-Cr and LR-NP-Cr). Ferromagnetic
nanoparticles (NP and NP-Cr) were also analyzed.
According to the findings in Fig.1, the prominent
bands of functional groups prior to and after sorption are
depicted. The prominent peaks in 3400, 2900, 1700, 1250,
and 1050 cm−1 correspond, respectively, to the asymmetric
and symmetric axial deformation of O–H, C–H, C = O, C–O,
and C–O–C, mainly attributed to the presence of cellulose
and lignin in both LR and LR-NP materials (Silverstein etal.
2012; Milani etal. 2018a).
The infrared spectra of LR and LR-NP prior to and after
Cr(VI) sorption exhibited prominent bands around 3200 to
3600 cm−1 due to the asymmetric and symmetric axial defor-
mation of alcohols OH, probably from cellulose, hemicellu-
lose, and lignin in the LR biomass, which in the lettuce roots
are more evident in this band between 3408 to 3424 cm−1
(Mothé and De Miranda 2009). After that, a band between
1722 and 1734 cm−1 is observed due to carbonyl groups
(C = O) from hemicellulose present in LR. The presence of
lignin and cellulose is evident by the asymmetric axial defor-
mation, representing methylene groups (CH2) and aliphatic
and aromatic C–H with symmetry verified at 2927-cm−1
absorption (Mothé and De Miranda 2009).
Bands between 1100 and 1200 cm−1 are related to hemi-
cellulose and cellulose in LR and LR-NP structures. The
Fig. 1 FTIR spectra of the biomass, nanocomposites, and magnetite
in study: lettuce root biomass (LR) prior to and (LR-Cr) after chro-
mium sorption; nanomodified lettuce root biomass (LR-NP) prior to
and after (LR-NP-Cr) chromium sorption; magnetite nanoparticles
(NPs) prior to and after (NP-Cr) chromium sorption. The spectra are
shifted relative to y-axis for better viewing

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presence of the band deformation in 1024 to 1040 cm−1 is
due to the bonds of the C–O–C groups of aliphatic and aro-
matic esters found in the biosorbent (Binod etal. 2012).
The band stretching at 1638 to 1645 cm−1 after chromium
adsorption indicates that aldehydes, esters, and carboxylic
acids are involved in the biosorption process (Babu etal.
2018). In addition, some bands were intensified after sorp-
tion, possibly related to the availability of Cr(VI) binding
sites about 3400 cm−1.
Bands attributed to magnetite are observed at 593 cm−1 in
the NP infrared spectrum. As predicted, after sorption, the
band shifted to 608 cm−1 due to the interaction of magnetite
with the lignin present in the root biomass, indicating the
complete formation and efficient impregnation of NP in LR
(Panneerselvam etal. 2011; José etal. 2019).
The micrographs shown in Fig.2 demonstrate the bio-
mass surface singularities before impregnation and after the
nanocomposite formation. According to the literature, the
irregular surface observed in LR (2A and 2B) and LR-Cr
(2C and 2D) is composed of polysaccharides, pectin, cel-
lulose, and hemicellulose, which contain carboxylic, citric,
and malic acids capable of interacting with metallic ions
in the aqueous medium (Akhter etal. 2014; Milani etal.
2018b).
The presence of magnetite nanoparticles can be observed
on LR surfaces in Fig.2 before (Fig.2E, F) and after
(Fig.2G, H) Cr(VI) sorption. It is also noticed that this mate-
rial exhibits non-uniform particle sizes commonly produced
when the coprecipitation method is used and some conglom-
erates of nanoparticles. As for the nanocomposite SEM, it is
possible to observe the root surface covered by Fe3O4 prior
(Fig.2I, J) and after (Fig.2K, L) Cr(VI) sorption. The same
fibrous morphology existing in LR and LR-Cr (Fig.2–D) is
also observed, however, with NP’s deposition on its surface
as indicated in Fig.2 I–L.
Thus, the images obtained in the morphological analysis
of in natura and nanomodified lettuce roots (Fig.2A, B, I, J)
show a rough and fibrous aspect with a wide heterogeneous
surface area. That favors the use of this material as an ideal
biosorbent for removing metal ions from an aqueous medium
(Schwantes etal. 2015). In the micrographs obtained to detect
the chemical distribution after Cr(VI) sorption (Fig.2C, D,
K, L), it is possible to observe small regions significantly
brighter. This observation may imply the occurrence of
greater electron backscattering, which coincides with the
interaction between the electron beam and heavy metals,
such as chromium. This contrast might denote the possible
presence of chromium ions on the surface of the biosorb-
ent, being identified when compared with the morphological
images before sorption.
In Fig.3, the diffractograms of nanoparticles and the
nanomodified material prior (NP and LR-NP) and after
(NP-Cr(VI) and LR-NP-Cr(VI)) sorption are compared to
demonstrate the efficiency of the biomass nanomodification
with magnetite. As for LR and LR-Cr(VI), no reflection
peaks are observed in the diffractograms since LR biomass
is composed of organic matter and exhibits an amorphous
structure.
Magnetite is chemically composed of ferrous and fer-
ric iron. It is described as iron oxide II and III, and its
structure shows an inverted spinel with alternating octa-
hedral and tetrahedral layers. It is possible to observe in
its structure that the ferrous species occupies half of the
octahedral sites for the possible stabilization of the metal
and the ferric species occupy the other half. In addition
to this structure, the magnetite unit cell adheres to the
face-centered cubic with a = 0.8396nm crystal structure
parameter (Blaney 2007). These particularities of the mag-
netite make it possible to carry out its characterization,
generating characteristic peaks due to its inverse spinel
structure, the confirmation of biomass impregnation, and
the nanoparticle in an efficient manner. The reflection
peaks with greater intensities correspond to the planes
2θ = 30.16° (220), 35.5° (311), 43.0° (400), 56.9° (511),
and 62.6° (440) (Labuto etal. 2018; Milani etal. 2018a;
José etal. 2019; Blaney 2007).
The LR-NP magnetic property observed by the presence
of the magnetite crystalline phases and verified by the char-
acteristics of Fe–O stretches in the FTIR analysis is also
demonstrated in the images of Fig.4. This illustration shows
the biomass magnetization strength as a neodymium magnet
approximates the LR-NP powdered material (Fig.4B) or an
LR-NP-Cr suspension (Fig.4D). It can be observed that the
magnet electromagnetically attracts the nanocomposite to it.
This process occurs at about 30s.
pH assessment: point ofzero charge andsorption
pH
At the pHPZC, the surface charge of the adsorbents is electri-
cally zero. When exposed to a solution with a pH below the
point of zero charge, the material will be positively charged.
If the adsorption media is at a pH greater than pHPZC, the
surface will be negatively charged (Fernández-Nieves etal.
1998; Zheng etal. 2009; Beretta etal. 2021). Figure5
depicts the graphical representation of the pHPZC of the
adsorbents LR (5.96) and LR-NP (6.50).
If the pH of aqueous solutions is below pHPZC, the mate-
rial surface is positively charged, showing a cationic form,
indicating that the sorption of anionic analytes is favora-
ble. The inverse behavior occurs for the adsorbent material
exposed to a value higher than its pHPZC. The material is
loaded negatively, indicating an interaction for the sorption
of cationic analytes (Appel etal. 2003; Furlan etal. 2010;
Gabriel etal. 2021). In aqueous solutions, when exposed to

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a pH below pHPZC, Cr(VI) ions are attracted to the surface
of lettuce roots and adsorbed as HCrO4−.
After determining the pHPZC, the pH effect on the sorp-
tion was performed in pH ranging from 1 to 6. The results
are presented in Fig.6. The results indicated that Cr(VI)
sorption by LR and LR-NP was favorable below the pHPZC,
where the biosorbent surface is positively charged. Above it,
the values decreased dramatically. At pH 1.0, both showed
the highest efficiency for chromium removal (approximately
86.7% for LR-NP and 87.6% for LR). This behavior was
due to the attraction of ions by positively charged functional
groups on the surface of the biosorbent.
In aquatic environments, hexavalent chromium can form
various species such as
CrO4
2−
,
HCrO4
−
, and
H2CrO4
,
depending on the solution’s analyte concentration, pH, and
the presence of oxidizing and reducing agents. The predomi-
nant chromium species at each pH are presented in Reactions
3 and 4 (Jobby etal. 2018; Omer etal. 2019; Liu etal. 2017).
The main form of hexavalent chromium in water
is the oxyanion. In lower pH values, positive charge is
due to functional groups such as amine and hydroxyl
protonated at low pH, demonstrating that the strength
(Reaction 3)
H2CrO4
→
H++ HCrO4
−(pH =
1
−
6
)
(Reaction 4)
HCrO4
−
→
H++ CrO4
2−(pH
>7
)
Fig. 2 Scanning electron microscopy (SEM) images of in natura let-
tuce roots (LR) A, B prior to and C, D after Cr(VI) sorption; magnet-
ite nanoparticles (NP) E, F prior to and G, H after Cr(VI) sorption;
nanomodified LR (LR-NP) I, J prior to and K, L after Cr(VI) sorp-
tion. SEM were obtained at 2K × and 10K × magnifications
◂
Fig. 3 X-ray diffractograms by magnetic nanoparticle (NP), in natura (LR) and nanomodified (LR-NP) lettuce roots prior to and after Cr(VI)
sorption

Environmental Science and Pollution Research
1 3
between biomass and the analyte was electrostatic.
With the increase in pH, it is observed that adsorption
decreases; this consequence is that the degree of protona-
tion decreases and can change to, for example, decreasing
electrostatic attraction (Li etal. 2020). This process is
described in two ways, the first is through the reduction of
Cr(VI) to Cr(III) during the oxidation of organic matter,
whereas the second mechanism suggests that the reduction
is after adsorption by the surface (Gonzalez etal. 2008;
Kratochvil etal. 1998; Yao etal. 2010).
Kinetic studies
The kinetic study results, which indicate the removal of
Cr(VI) as a function of the contact time between adsorbent
and contaminant at a previously established pH (1.0), are
shown in Fig.7. It is possible to verify that in the first 5min,
chromium retention occurs and that the equilibrium was
reached in 30min for both biosorbents (LR and LR-NP),
demonstrating the advantage to use these materials in flow
systems for the treatment of effluents.
Table1 and Figure S2 (Supplementary Material)
exhibit the results of the pseudo-first- and pseudo-second-
order kinetics models applied to the experimental data.
It is observed that for both materials, LR (r2 = 0.9982)
and LR-NP (r2 = 0.9812), the pseudo-second-order equa-
tion provided the best fit since their r2 were close to 1.
In addition, this kinetic model was adequate to describe
the sorption phenomenon because the Qe values obtained,
0.523mg/g for LR and 0.405mg/g for LR-NP (Table2),
were comparable to those found experimentally for LR
Fig. 4 Illustration of A powdered LR-NP nanocomposite and B its
magnetization effect; LR-NP in Cr(VI) solution C prior to and D after
approximation of a neodymium magnet to this suspension
Fig. 5 Point of zero charge (pHPZC) using 1.0g/L dosage of A in nat-
ura (LR) and B nanomodified (LR-NP) lettuce roots with 0.1mol/L
NaCl solution at 2–12 pH range
Fig. 6 Effect of pH on Cr(VI) sorption capacity by in natura (LR)
and nanomodified (LR-NP) lettuce roots, at 10g/L adsorbent dosage
and 10mg/L Cr(VI) solution. n = 3

Environmental Science and Pollution Research
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(0.516mg/g) and LR-NP (0.403mg/g). The r2 found for
the pseudo-first order model was, respectively, 0.3608
and 0.6918 for LR and LR-NP, and did not describe the
data satisfactorily. Thus, the biosorption process was bet-
ter described by the pseudo-second-order model, which
corroborates the assumption that a chemisorption process
takes place (Babu etal. 2019).
Sorption capacity andisotherm model parameters
The batch procedure (Fig. S1, Supplementary Material)
was used to determine the sorption capacity of the biosor-
bents. Increasing concentrations of 10 to 100 Cr(VI) mg/L,
using magnetite NPs, and LR and LR-NP. All results of this
approach are presented in Table2 and illustrated in Fig.8.
The sorption capacity reached for LR, NP, and LR-NP was
4.51 ± 0.04mg/g, 2.48 ± 0.57mg/g, and 3.84 ± 0.08mg/g,
respectively. Although nanomodified biomass presented
slightly higher Qexp, nanomodification is desirable to pro-
vide easier removal of the biosorbent from the medium by
using a magnet to attract the material with superparamag-
netic properties.
In addition to biomasses, other materials also have sites
available to remove contaminants. Thus, adsorption by iron
oxide (NP) occurs mainly due to the hydroxyl groups pre-
sent on its surface, which tend to vary at different pH values
(Hu etal. 2004). Thus, anion adsorption is favored below
the point of zero charge of the adsorbent material. At the
same time, at higher pHs, there is competition for the sorp-
tion sites between hydroxyl ions and Cr(VI) species in the
form of chromate. In addition, this difference in the sorption
capacity can be attributed to the affinity of magnetite with
the different species that coexist at more acidic pHs (Hu
etal. 2004).
According to the Giles etal. (1974) classification, the iso-
therms presented in Fig.8 are type L, indicating a promising
process and demonstrating affinity between adsorbent and
adsorbate (Volesky 2004; Limousin etal. 2007). In Table2,
considering the algorithm χ2 (Labuto etal.2018), it is noted
that the presented theoretical models can be satisfactorily
applied to the experimental data of Cr(VI) sorption by all
adsorbents tested (LR, LR-NP, and NP).
When analyzing the r2, it was observed that Langmuir,
Freundlich, and Sips models presented the best values
(0.9628, 0.9580, and 0.9577, respectively) for LR. On the
other hand, it is observed that the Qmax values provided by
Langmuir (9.14 ± 1.6mg/g) are considerably higher than
those found experimentally (4.51 ± 0.04mg/g). Besides
presenting higher Qmax than expected (10.29 ± 8.28mg/g
and 9.14 ± 5.75mg/g, respectively), Sips and Hill models
showed unacceptable SE values.
Although the Freundlich model does not predict adsor-
bent saturation, other parameters, such as nf, help understand
the process since when it presents values between 1 and 10,
it indicates that the process occurred favorably (José etal.
2019). In this case, the process was favorable, not only for
LR (nf = 1.55) but for the other materials (1.38 and 1.80 for
LR-NP and NP, respectively).
When analyzing the D-R model for LR, it is noted that Qmax
provided by the model was fairly similar to Qexp. However,
low r2 values were found among all the applied models. D-R
isotherm provides parameters for calculating free energy (E),
allowing a better understanding of the process type: chem-
osorption and physisorption. For all materials, energy (E)
values were low (0.11, 0.10, and 0.08kJ/mol for NP, LR, and
LR-NP, respectively), indicating that sorption exhibits a physi-
cal nature as demonstrated by the adjustment of Freundlich’s
Fig. 7 Cr(VI) sorption kinetics employing 5, 10, 15, 20, 25, 30, 60,
90, 120, and 150min contact time, 10 g/L dosage of in natura (LR)
and nanomodified lettuce roots composite (LR-NP), and 10 mg/L
Cr(VI) solution at pH 1.0. n = 3
Table 1 Information of pseudo-first-order kinetics and pseudo-second-order by LR (in natura lettuce root biomass), and LR-NP (nanomodified
lettuce root biomass), using 100mg of biosorbent suspended in 10mL of 100mg/L Cr(VI) solution. n = 3
Pseudo first order Pseudo second order
r2K1 (min−1)qexp (mg/g) r2K2 (g/mg min) qexp (mg/g)
LR 0.3608 0.084 0.1441 0.9982 0.3724 0.5238
LR-NP 0.6918 0.091 0.1876 0.9812 3.77 × 1017 0.4059

Environmental Science and Pollution Research
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model. It considers a physical process in multilayers and het-
erogeneous surfaces (Foo and Hameed 2010).
For the LR-NP, the same isothermal models (Langmuir,
Freundlich, and Sips) obtained better adjustments consider-
ing r2 and χ2. However, the Qmax value provided by Lang-
muir (8.63 ± 1.05mg/g) was also higher than the experimen-
tal (3.84 ± 0.08mg/g) while Sips, in addition to Qmax, also
showed a high SE value (10.11 ± 6.63mg/g).
On the other hand, the Freundlich model indicated that
the sorption of Cr(VI) by LR-NP was favorable and with
r2 = 0.9909 and χ2 = 0.0165, presenting an excellent adjust-
ment to the experimental data and indicating that the process
is physical, as well as for the LR. It is worth mentioning that
the D-R model presented acceptable values of r2 (0.9076)
and χ2 (0.1687). However, they were not the best, demon-
strating to fit well with the experimental data, given the
proximity of the Qmax value (3.70 ± 0.3mg/g) to that found
experimentally (3.84 ± 0.08mg/g). Thus, free energy E, cal-
culated by the D-R model, can also indicate that the process
is physisorption.
As for NP, the D-R model provides the best r2 value
(0.9471), in addition to Qmax (2.38 ± 0.12mg/g) being very
close to the experimental value found (2.48 ± 0.57mg/g).
Considering that the free energy supplied from the BDR
parameter is low (0.11kJ/mol), the process is physical, as
suggested for LR and LR-NP. In this way, the Freundlich
Table 2 Data of experimental
sorption capacity (Qexp),
isotherms parameters and
χ.2 error evaluation for
Cr(VI) sorption by magnetic
nanoparticles (NP), in natura
(LR), and nanomodified (LR-
NP) lettuce roots. SD standard
deviation, SE standard error
provided by fitting the model to
the experimental data. n = 3
NP LR LR-NP
Qexp (mg/g) 2.48 ± 0.57(SD) 4.51 ± 0.04(SD) 3.84 ± 0.08(SD)
Isotherm model
Langmuir
Qmax (mg/g) 3.77 ± 0.7(SE) 9.14 ± 1.8(SE) 8.63 ± 1.054(SE)
b (L/g) 0.0229 ± 0.0093(SE) 0.0164 ± 0.0056(SE) 0.0098 ± 0.0019(SE)
r20.9185 0.9628 0.9925
χ20.0721 0.1066 0.0137
Freundlich
Kf
(L/mg) 0.2185 ± 0.0971(SE) 0.3249 ± 0.0445(SE) 0.16 ± 0.02(SE)
nf
1.80 ± 0.36(SE) 1.55 ± 0.18(SE) 1.38 ± 0.07(SE)
r20.8876 0.9580 0.9909
χ20.0994 0.1202 0.0165
D-R
QDR
(mg/g) 2.38 ± 0.12(SE) 4.62 ± 0.44(SE) 3.70 ± 0.3(SE)
BDR(mol2∕J2
)3.91 × 10−5 ± 7.7 × 10−6(SE) 4.67 × 10−5 ± 1.64 × 10−5(SE) 7.39 × 10−5 ± 1.89 × 10−5(SE)
E (kJ/mol) 0.11 0.103 0.08
r20.9471 0.8498 0.9077
χ20.0468 0.43 0.1687
Sips
Qmax (mg/g) 2.47 ± 0.25(SE) 10.3 ± 8.28(SE) 10.1 ± 6.63(SE)
Ks0.050 ± 0.0084(SE) 0.013 ± 0.021(SE) 0.007 ± 0.0088(SE)
n2.12 ± 0.70(SE) 0.93 ± 0.34(SE) 0.94 ± 0.20(SE)
r20.9352 0.9577 0.9914
χ20.0574 0.1211 0.0157
Temkin
bT9.8 × 104 ± 1.0 × 104(SE) 6.2 × 104 ± 7.5 × 103(SE) 7.1 × 104 ± 6.8 × 103(SE)
K (L/mg) 0.18 ± 0.04(SE) 0.30 ± 0.08(SE) 0.19 ± 0.04(SE)
r20.9450 0.9194 0.9506
χ20.0487 0.2308 0.0901
Hill
QH (mg/g) 2.47 ± 0.25(SE) 9.14 ± 5.75(SE) 9.34 × 109 ± – (SE)
nH2.12 ± 0.73(SE) 1 ± 0.35(SE) 0.72 ± 0.041(SE)
KH565.6 ± 1152.4(SE) 60.95 ± 27.42(SE) 5.747 × 1010 ± – (SE)
r20.9352 0.9575 0.9895
χ20.0574 0.1218 0.0193

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1 3
model can also be adopted to describe the experimental data
of Cr(VI) adsorption by NP.
Conclusions
Chromium(VI) sorption by LR and LR-NP proved to be
more efficient with decreasing pH, reaching maximum
efficiency at pH 1. From the experimental data of sorp-
tion kinetics, it was noticed that the process occurs as
early as the first 5min and that equilibrium was reached
after 30min of contact time. The pseudo-second-order
model showed a better fit to the experimental data, sug-
gesting that a chemosorption process occurs. Usingthe
previous set parameters pH and kinetics, the sorp-
tion capacity studies were performed in 30mincon-
tact time,pH 1.0, and 10 g/L LR and LR-NPdose, so
that it was possible to compare the effect of magneti-
zation. Experimental sorption capacity was 4.51mg/g
and 3.84mg/g for LR and LR-NP, respectively, dem-
onstrating that nanomodification besides facilitating the
removal of the metal ion did not reduce sorption capac-
ity significantly. Among the isotherm models applied to
experimental data, Freundlich was the model that best
described the sorption process byLR, NP, and LR-NP.
Therefore, the adsorbentmaterials proposedin this work
areefficient and low-cost alternatives for removing
Cr(VI) from water.
Supplementary Information The online version contains supplemen-
tary material available at https:// doi. org/ 10. 1007/ s11356- 022- 21755-0.
Author contribution EC and GL postulated and supervised the study.
BCS and EC planned the experiment. BCS and TEA obtained the data,
and EC, GL, and JCJ carried out the data analysis and interpretation.
GL performed the adjustment of all experimental data to the isothermal
models applied. BCS, JCJ, and TEA prepared the first draft, and EC
and GL thoroughly revised the manuscript. EC, GL, and BCS read and
approved the final manuscript.
Funding This work was supported by Fundação de Amparo à Pesquisa
do Estado de São Paulo (FAPESP) (2016/06271–4 and 2019/08335–8)
and Conselho Nacional de Desenvolvimento Científico e Tecnológico
(CNPq) (163205/2019–1 and 166783/2019–6).
Availability of data and materials All data generated or analyzed during
this study are included in this published article. Extra Data are available
from the authors (elma.carrilho@gmail.com) upon reasonable request.
Declarations
Ethical approval Not applicable.
Consent to participate Not applicable.
Consent to publish Not applicable.
Competing interests The authors declare no competing interests.
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