BY-NC-ND 3.0 license Open Access Published by De Gruyter March 24, 2018

p-Nitrophenol determination and remediation: an overview

Francis Merlin Melataguia Tchieno and Ignas Kenfack Tonle


An almost exhaustive overview of the determination and remediation of p-nitrophenol (p-NP)-contaminated media is herein presented. p-NP is one of the priority pollutants on the U.S. Environmental Protection Agency list. This is because p-NP is either a precursor or a derivative of a good number of pollutants. It is itself very carcinogenic and tends to persist in water and soil. This has prompted the development of a wide range of analytical tools by researchers for its determination and eventual removal from contaminated sites. These include electrochemical methods with many electrode modifiers and electroanalytical procedures developed for the detection/quantification of p-NP in contaminated environments. Also, chromatographic and p-NP uptake techniques, particularly adsorption onto various adsorbents (ranging from natural to synthetic), are reviewed. The use of microorganisms for the bioremediation of p-NP-contaminated samples and sites has equally been largely studied and is herein overviewed, not forgetting advanced oxidative processes.


Bioaccumulation and biomagnification of xenobiotics in aquatic and terrestrial ecosystems is on the rise due to their massive mobilization in an attempt to satisfy the ever-rising needs of mankind. Among these xenobiotics are the nitroaromatics, particularly nitrobenzene, dinitrotoluenes, and mono- and di-nitrophenols, which are powerful carcinogens (Boehncke et al., 2000; Nishino, Spain & He, 2000). Their presence (above 0.7 mm) in irrigation water leads to the reduction of crop output (Luz et al. 2004). Nitroaromatics and nitrophenols, in particular, take part in the synthesis of many industrial products such as pesticides, herbicides, petrochemicals, explosives, pharmaceuticals, synthetic dyes (Kulkarni & Chaudhari, 2007; Cushing et al., 2012; Wang et al., 2012), and processing of leather (Jiao, Luo, and Li 2013). Nitrophenols appear in the degradation of pesticides such as parathion and nitrofen (Di Paola et al. 2003). For these reasons, nitroaromatics have been blacklisted by the U.S. Environmental Protection Agency (EPA) as high-priority toxic pollutants (EPA and 2008). p-Nitrophenol (p-NP) is one of these recalcitrant nitroaromatics, owing to its high stability and solubility. Its detection/determination and removal from contaminated sites is therefore of prime importance.

Various analytical techniques have been used for the determination of p-NP and/or its isomers, among which fluorescence detection, high-performance liquid chromatography (HPLC), spectrophotometry, capillary electrophoresis, and electrochemical methods (Tang et al. 2015a). Chromatographic techniques are quite excellent from an analytical point of view, but practically they require expensive, sophisticated, and heavy equipment. Additionally, they need highly trained operators and often-tedious sample pretreatment (Tonle et al. 2015). Electrochemical methods are a convenient alternative to these chromatographic methods for the detection/determination of p-NP because they require simple operational procedures and low-cost equipment, they are highly accurate and sensitive, and experimental data are obtained in real time (Zhang et al., 2014a; Tonle et al., 2015; Tang et al., 2015a). For even better p-NP detection limits, working electrode surfaces have been modified (El Mhammedi et al., 2009; Liu et al., 2009; Xu et al., 2013; Wei et al., 2015; Ahmad et al., 2016).

Concerning p-NP remediation, various approaches have been used by researchers, among which adsorption onto chosen substrates, solvent extraction, advanced oxidative processes, microwave irradiation, chemical oxidation or reduction, and bioremediation (Longli et al., 2005; Kulkarni & Chaudhari, 2007). Adsorption and solvent extraction only transfer the p-NP from one phase to another but do not destroy it and a secondary treatment is required. Chemical oxidation and reduction is not only costly (severe reaction conditions-high temperature and pressure) but also could lead to low mineralization (Longli et al. 2005). Advanced oxidative processes, on their part, could generate toxic intermediates and involve enormous cost (Kanekar, Doudpure, and Sarnaik 2003). Bioremediation is by far the most cost-effective and eco-friendly approach for the removal of p-NP from contaminated sites. Microbes are endowed with inherent abilities to live, metabolize, thrive, colonize, and decontaminate p-NP-contaminated media (Kulkarni and Chaudhari 2007). These include species from the genera Arthrobacter, Bacillus, Flavobacterium, Moraxella, Nocardia strain, Pseudomonas, and Rhodococcus, among others (Leung et al. 1997). With the appropriate microbial strain chosen, complete mineralization of p-NP can be obtained.

p-NP determination techniques

Electrochemical methods

Electrochemical methods are especially suitable for the large-scale environmental monitoring of electrochemically active species because they are inexpensive, extremely sensitive, and applicable as an independent alternative to so far prevalent spectrometric and separation techniques (Stanic and Girousi 2011). They can give a large useful concentration range for both inorganic and organic species (Svancara et al. 2009). Of great interest in electroanalysis are devices based on modified electrodes that can be easily designed for the sensitive and selective detection of a given analyte. Among all the chemical sensors reported in the literature, electrochemical sensors are the most attractive because of their remarkable sensitivity, experimental simplicity, and low cost (Mousty 2004).

Among the electrochemical methods, cyclic voltammetry (CV) is the most straightforward and most widely used for the investigation of the mechanisms of redox systems. Zunic et al. (2014) used an organically modified bentonite (BT) to construct a clay-based p-NP sensor. The CV response of the intercalated clay film at the surface of glassy carbon electrode (GCE) increased with the increasing concentration of p-NP in solution. The detection limit of p-NP at the benzyltrimethylammonium (BTMA)-BT/GCE was 7 μm. Using a similar organoclay-modified GCE configuration and with the aid of CV, p-NP was detected in the linear concentration range 2–16 mm (Rabi-Stankovic et al. 2012). Pontié et al. (2011) successfully immobilized a nickel tetrasulfonated phthalocyanine (p-NiTSPc) membrane on GCE, which they used to oxidize p-NP with limited electrode passivation. The lack of passivation was due to the formation of water-soluble p-aminophenol intermediate at low potentials. The polymer film-modified GCE was twofold more sensitive than bare GCE and a linear concentration range of 20–100 mg/l was obtained using CV. The CV response obtained at bare GCE is shown in Figure 1. Peaks 1–3 were assigned to the oxidation/reduction of the nitro function, whereas peak 4 was attributed to the oxidation of the phenol group. Liu et al. (2009) designed a nanoporous Au electrode for the investigation of the redox behavior of p-NP by CV. With high sensitivity and good selectivity of the nanoporous Au samples, the authors could detect p-NP in the linear range from 0.25 to 10 mg/l.

Figure 1: CV of 50 mg/l p-NP in acetate buffer (pH 5.2, ionic strength 0.2 m), on a polished GCE, under a potential scan rate of 100 mV/s (Pontié et al. 2011).© Copyright Wiley-VCH Verlag GmbH & Co. and reproduced with permission.

Figure 1:

CV of 50 mg/l p-NP in acetate buffer (pH 5.2, ionic strength 0.2 m), on a polished GCE, under a potential scan rate of 100 mV/s (Pontié et al. 2011).

© Copyright Wiley-VCH Verlag GmbH & Co. and reproduced with permission.

Various authors have proposed different mechanisms for the oxidation of p-NP at electrode surfaces. These pathways are shown in Figure 2. In mechanism III, a p-aminophenol film is first formed at the surface of the electrode (due to the low starting potential) and oxidized to give benzoquinone as by-product (Barbero, Silber & Sereno, 1989; Pontié et al., 2011).

Figure 2: Possible p-NP electrochemical oxidation pathways: mechanism I (Zunic et al. 2014; © Copyright Elsevier BV), mechanism II (Liu et al. 2009; © Copyright Elsevier BV), and mechanism III ().Reproduced with permission.

Figure 2:

Possible p-NP electrochemical oxidation pathways: mechanism I (Zunic et al. 2014; © Copyright Elsevier BV), mechanism II (Liu et al. 2009; © Copyright Elsevier BV), and mechanism III ().

Reproduced with permission.

Using different electrode configurations, several authors successfully employed linear sweep voltammetry (LSV) in the electrochemical detection of p-NP with varying degrees of success. In this light, Tang et al. (2015a) proposed a graphene-chitosan film-modified GCE that could detect p-NP down to 0.09 μm by LSV. This electrode configuration was successfully applied to the detection of p-NP in lake water, river water, and wastewater without any pretreatment. With the same electrochemical method, p-NP was detected down to 0.16 μm on an activated carbon (AC)-modified GCE by Madhu et al. (2014). Other contributions employing LSV unmodified and modified electrodes for the sensing of p-NP are summarized in Table 1.

Table 1:

Summary of the determination of p-NP at unmodified and modified electrodes.

Electrode configurationElectrochemical methodLinear range (μM)Detection limit (μM)Reference
Nanoporous AuCV17.97–71.89Liu et al. 2009
p-NiTSPc/GCECV143.77–718.86Pontié et al. 2011
BTMA-BT/GCECV10–1007Zunic et al. 2014
α-MnO2 nanotube/GCECV100–700100Wu et al. 2014
ZnO/GCECV0.01–100013Sinhamahapatra, Bhattacharjya, and Yu 2015
BTMA-BT/GCECV2000–16,000Rabi-Stankovic et al. 2012
Oxide/PSi/PThCV0.015–300.006Belhousse et al. 2014
Au disk electrodesCVForryan et al. 2004
GO/GCELSV0.1–1200.02Li et al. 2012
MWCNT/GCELSV2–40000.4Luo et al. 2008
AC/GCELSV1–5000.16Madhu et al. 2014
Nanoporous AuLSV28.75–71.860.144Liu et al. 2011
PtLSV10–6006.0Khachatryan et al. 2005
SPCELSV1–1001.0Fanjul-Bolado, González-García, and Costa-García 2006
Ag/MWCNT/GCELSV3.0–1201.3Liu et al. 2014a, 2014b
GR-CS/GCELSV0.1–1400.09Tang et al. 2015a
β-CD-SBA/CPEDPV0.2–1.60.01Xu et al. 2011
HMDEDPV0.72–14.380.017Ni, Wang, and Kokot 2001
Nafion/GCEDPV20–23017.1Calvo-Marzal et al. 2001
Polyfurfural film/GCEDPV0.75–1000.04Wei et al. 2015
OMCs/GCEDPV2–900.1Zhang et al. 2013a
SWy-2-AQ/GCEDPV0.22–323.490.144Hu et al. 2001
LiTCNE/PLL/GCEDPV0.27–23.2000.0075Luz et al. 2004
β-CD-RGO/GCEDPV7.19–71.890.359Liu et al. 2012
GCEDPV2–1000.17 (CR)Pfeifer et al. 2015
0.39 (AO)
GCE-RGO/SrTiO3DPV0.3–0.80.11Ahmad et al. 2016
p-AgSAEDPV2–102.8Fischer et al. 2007
Mg(Ni)FeO/CPEDPV2–2000.20Xu et al. 2013
MWCNT-Nafion/GCEDPV0.1–1.00.04Huang, Yang, and Zhang 2003
SWCNT/GCEDPV0.01–5.00.0025Yang 2004
Nano-Cu2O/PtDPV10–10000.10Gu et al. 2010
Ag particles/GCEDPV1.5–1400.50Casella and Contursi 2007
DTD/Ag/CPEDPV1–1000.25Rounaghi, Kakhki, and Azizi-Toupkanloo 2012
AuNP/RGO/GCEDPV0.05–2.00.01Tang et al. 2013a; 2013b
c/p-NiTSPc/CFMESWV0.08–71.890.288Tapsoba et al. 2009
HAP-CPESWV0.2–1000.008El Mhammedi et al. 2009
TMA-BT/GCESWV0.2–2001.0Rabi-Stankovic et al. 2013
AgSAESWV0.551–4.860.0519De Souza, Mascaro, and Fatibello-Filho 2011
BDDESWV30–500.03 (CR)Pedrosa, Codognoto, and Avaca 2003
0.02 (AO)
BiFE/GCESWV (BM)0.02–35.940.01Hutton, Ogorevc, and Smyth 2004
CAD (FI)CAD (FI)0.004
Poly(p-aminobenzene sulfonic acid)/GESDV3–7000.3Yao et al. 2015
AuNP/GCESDV10–10008.0Chu, Han, and Zhang 2011
EGECEMPA200–100050–100Bebeselea et al. 2008
AuNP-SPCEHC0.1–3150.098Tang and Chen 2011
MWCNT-PDPA/GCEHV8.9–1430Yang, Unnikrishnan, and Chen 2011
DMEDPP0.5–100Asadpour-Zeynali and Soheili-Azad 2012
G-Au/GCEChronoamperometry470–10,7500.47Zhang et al. 2012a
[PSS-RGO/PAMAM-AuNPs]20/AuEChronoamperometry5–5151.8Huang et al. 2015
AgSAEHPLC-ED10–250010 (TLD)Danhel et al. 2009
25–250025 (WJD)
PCEHPLC-ED0.004–0.1441.01Galeano-Díaz et al. 2000
Poly(Ni-(PPIX)]/GCEHPLC-EDUp to 9.350.093Dall’Orto et al. 1996
GCEHPLC-ED7–500 ng on column1 ng on columnHoneychurch and Hart 2007
CuO nanocubes/GCEI-V technique0.01–10000.005Abaker et al. 2012
SWCNH/GCELSV0.05–100.011Zhu et al. 2009
PB/PtSWV10–9010.28Lupu et al. 2009

  1. AgSAE, Ag solid amalgam electrode; APTEMS, amino-propyltrimethoxysilane; AuE, Au electrode; AuNP, Au nanoparticle; BDDE, boron-doped diamond electrode; BiFE, bismuth film electrodeposited; BM, batch mode; c/p-NiTSPc, coated p-NiTSPc; CFME, carbon fiber microelectrode; DME, dropping mercury electrode; DTD, 6,7,9,10,17,18,19,20,21,22-decahydrodibenzo[h,r][1,4,7,11,15]trioxadiazacyclonanodecine-16,23-dione; EGECE, expanded graphite-epoxy composite electrode; FI, flow injection; G-Au, Au nanoparticles decorated graphene sheet; GE, graphite electrode; GR-CS, graphene-chitosan; HAP, hydroxyapatite; LiTCNE, lithium tetracyanoethylenide; m-AgSAE, mercury meniscus-modified AgSAE; MWCNT, multiwalled CNT; OMCs, ordered mesoporous carbons; oxide/PSi, oxidized hydrogen-terminated porous silicon; p-AgSAE, liquid mercury-free polished AgSAE; p-PPD, p-phenylenediamine; PAMAM, polyamidoamine; PCE, porous carbon electrode; PDPA, poly(diphenylamine); PLL, poly-L-lysine; PPIX, protoporphyrin IX; PSS, poly sodium 4-styrenesulfonate; PTh, polythiophene; SBA, mesoporous silica SBA-15; SPCE, screen printing carbon electrode; STV, steady-state voltammetry; TLD, thin-layer detector; WJD, wall-jet detector; PAVT, poly(3-[(E)-2-azulene-1-yl)vinyl]thiophene; PB, Prussian blue; PAZ, polyazulene.

A number of electrode configurations have also been proposed for the electroanalytical detection of p-NP by differential pulse voltammetry (DPV) as shown in Table 1. These range from film-modified electrodes (Calvo-Marzal et al., 2001; Luz et al., 2004; Asadpour-Zeynali & Najafi-Marandi, 2011; Wei et al., 2015) through carbon paste electrodes (CPEs; Xu et al., 2011; Rounaghi, Kakhki & Azizi-Toupkanloo, 2012; Xu et al., 2013) to hanging mercury drop electrodes (HMDEs; Ni, Wang, and Kokot 2001). A perovskite/reduced graphene oxide (RGO) nanocomposite was fabricated and successfully immobilized at the surface of GCE (Ahmad et al. 2016). This electrode configuration gave a detection limit of 110 nm for p-NP using DPV and the authors could recover p-NP in spiked tap water up to 97%. Another film, polyfurfural film/GCE, prepared by Wei et al. (2015) gave a detection limit of 0.04 μm for p-NP.

Square-wave voltammetry (SWV), semiderivative voltammetry (SDV), cathodic amperometric detection (CAD), multiple pulse amperometry (MPA), hydrodynamic chronoamperometry (HC), differential pulse polarography (DPP), hydrodynamic voltammetry (HV), chronoamperometry, and HPLC with electrochemical detection (HPLC-ED) have also been employed for the sensing of p-NP (Table 1).

Chromatographic methods

p-NP has been identified and quantified by chromatographic techniques, particularly HPLC and gas chromatography coupled to other detection techniques such as nitrogen/phosphorus-sensitive detection, fluorescence detection, mass spectrometry, diode array detection, and electrochemical detection. Using HPLC with tandem mass spectrometry detection, Barr et al. (2002) developed a sensitive and selective method for the quantification of p-NP. They applied this method to the analysis of about 16,000 samples collected from residents exposed to methyl parathion and/or p-NP from their homes, which were contaminated by the illegal application of methyl parathion. A similar study was carried out by Hryhorczuk et al. (2002). Another approach based on supercritical fluid extraction and gas chromatography determination was developed by Wong et al. (1991) for the quantification of p-NP in soils.

With HPLC-based determination methods, varied detection limits for p-NP have been reported, among which 0.25 μg/l (Nick & Schöler, 1992; Neng & Nogueira, 2014), 1.0 μg/l (Han, Jung, and Shin 2008), 0.062 ng/μl (Belloli et al. 1999), 96 ng injected (Fiehn and Jekel 1997), 1 mg/l (Brega 1990), and 3.54 mg/l (Bigley and Grob 1985). The following limits of quantification were also reported for the quantification of p-NP using chromatographic methods coupled to other techniques: 2.5 μm (Almási, Fischer, and Perjési 2011) and 1 μm (Elbarbry, Wilby, and Alcorn 2006).

Other determination methods

Manera et al. (2007) reported the first combination of flow-through diffuse reflectance optosensing with multivariate regression modeling for the trace level determination of p-NP. The authors could use this method to determine p-NP down to 0.003 μm, without the use of hazardous solvents. In two other completely different approaches, Muresanu, Copolovici, and Pogacean (2005) and Guo, Wang, and Zhou (2004) were able to determine p-NP with rather good results. These methods together with their working ranges are summarized in Table 2.

Table 2:

Other p-NP determination methods.

MethodLinear range (μM)Detection limit (μM)Reference
Kinetic method9–253.0Muresanu, Copolovici, and Pogacean 2005
Diffuse reflectance optosensing0.007–0.60.003Manera et al. 2007
HPCZE20.3–40604.06Guo, Wang, and Zhou 2004

  1. HPCZE, high-performance capillary zone electrophoresis.

p-NP remediation techniques

Adsorption methods

Onto AC, char, and biochar

AC refers to a microporous adsorbent of carbonaceous origin prepared by heating carbon-containing materials to very high temperatures, causing burn-off of noncarbon impurities, in the absence of air and results in the formation of a char. This is followed by the oxidation of the char with oxidizing gases such as steam, air, or CO2 at high temperatures to impart activity and develop a porous structure with large surface area (Lynam, Kilduff, and Weber 1995). When the carbon-containing material is biomass, the char obtained is referred to as biochar (Zheng et al. 2017). Two samples of AC fibers and two samples of granulated ACs were used for the adsorption of p-NP from aqueous solution. The surface of these carbons was modified by oxidation with nitric acid and oxygen gas. The uptake of p-NP at the resulting carbon surfaces decreased after oxidation and increased after degassing. The presence of acidic surface groups suppressed the adsorption of p-NP, whereas the presence of nonacidic surface groups enhanced p-NP adsorption (Goyal 2004). p-NP adsorption has also been reported on Jatropha curcas AC (Azeez and Adekola 2016), unmodified and chemically modified commercial ACs (Terzyk et al. 2008), fiber-based AC from coconut husks (Al-Aoh et al. 2013), sludge-based AC (Mohamed et al. 2011), olive stone AC (Haydar et al. 2003), jute stick char (Ahmaruzzaman and Gayatri 2010a), AC derived from walnut peel (Liu, Wang, and Bai 2015), coal briquette char (Li, Roddick, and Hobday 1998), Fe/Zn biochar (Wang et al. 2017), and biochar from Chorella sp. Cha-01, Chlamydomonas sp. Tai-03, and Coelastrum sp. Pte-15 (Zheng et al. 2017). Coal fly ash (Alinnor and Nwachukwu 2011) and NaOH-modified palm oil fuel ash (Al-Aoh et al. 2012) have also been reported as efficient p-NP adsorbents. AC- and biochar-based materials for the uptake of p-NP are summarized in Table 3.

Table 3:

Adsorption of p-NP onto pristine and modified adsorbents.

AdsorbentEquilibrium timeQmax (mg/g)Reference
Olive cake-based AC120 minAbdel-Ghani, Rawash, and El-Chaghaby 2016
Acid-activated jute stick char4 h39.38Ahmaruzzaman and Gayatri 2010a
Activated tea waste 5 h142.85Ahmaruzzaman and Gayatri 2010b
Petroleum coke72 h11.06Ahmaruzzaman and Sharma 2005
Rice husk15.31
Rice husk char39.21
Residual coal86.95
Residual coal treated with H3PO4256.40
Coke breeze4.64
Palm-tree fruit stones120 min (303 K)136.62Ahmed and Theydan 2012
120 min (313 K)147.26
120 min (323 K)161.44
FeCl3 AC184.86Ahmed and Theydan 2015
NaOH-modified palm oil fuel ash5 h500Al-Aoh et al. 2012
AC fiber1 h500Al-Aoh et al. 2013
Granular AC3 h333.3
Fly ash2 h (30°C)26.5Alinnor and Nwachukwu 2011
2 h (40°C)19.8
2 h (60°C)11.5
F400<7 days244.69Alvarez et al. 2005
Chloromethylated styrene-divinylbenzene copolymer6 hArdelean, Davidescu, and Popa 2010
Olefin-grafted poly(styrene-co-divinylbenzene)24 h16.6Ardelean et al. 2012
Spectracarb 222524 h (H2O)292.13Ayranci and Duman 2005
24 h (1 m H2SO4)276.83
24 h (0.1 m NaOH)56.34
Activated kaolinitic clay120 min0.281Azeez and Adekola 2016
J. curcas AC1.82
CS 150112 h280.9Brasquet, Subrenat, and Cloirec 1999
RS 130112 h333.3
NC 6048 h301.2
HDTMA-palygorskite30 min137.74 Chang et al. 2009
MgAl-mixed oxide12 h367.8 Chen et al. 2009
MOF Cr-BDC24 h19 Chen et al. 2017
Granular AC4 days613.48Chern and Chien 2002
Granular AC4 daysChern and Chien 2003
p-DMAC168–12 min49.75Erdem et al. 2009
ACsGoyal 2004
Natural zeolite90 min0.54Guo and Wang 2016
Bagasse fly ash1.15 Gupta et al. 1998
Yallourn coal12.0Haghseresht and Lu 1998
Briquette char95.0
Tower char80.0
F1004 days (pH 2)237.88Haghseresht, Nouri, and Lu 2003
4 days (pH 12)65.38
HNO3-treated F1004 days (pH 2)161.37
4 days (pH 12)52.86
H2-treated F1004 days (pH 12)87.64
AC7 days500Haydar et al. 2003
Power station char7 days80Hobday et al. 1994
Briquette char50 h95
Yallourn coal12
Yallourn coarse grus14
Montmorillonite2 h15.30Houari et al. 2014
Hyper-cross-linked resin HJ-120 h (300 K)192.7Huang, Yan, and Huang 2009
20 h (305 K)189.8
20 h (310 K)183.5
XAD-420 h (300 K)43.54
20 h (305 K)46.06
20 h (310 K)45.19
Graphene24 h15.5Ismail 2015
Powdered AC2 hIvancev-Tumbas et al. 2008
Brazilian peat100 min (15°C)23.39Jaerger et al. 2015
100 min (25°C)22.90
100 min (35°C)16.10
Amino-MIL-53(Al)48 h297.85 Jia, Jiang, and Wu 2017
CPC-modified nano-SiO224 hJing et al. 2013
Cloisite-10A24 h21.70Ko et al. 2007
HDTMA-bentonite2 h (30°C)107.53Koyuncu et al. 2011
2 h (35°C)212.77
2 h (40°C)133.33
PEG-bentonite8 h (30°C)243.90
8 h (35°C)270.27
8 h (40°C)344.83
Activated FAS-1 charcoal10 days (293 K)388.12Krasil’nikova, Kul’kova, and Larin 2008
10 days (308 K)351.95
10 days (318 K)388.04
10 days (328 K)315.76
Granular AC48 h206.3Kumar et al. 2007
Salicylaldehyde functionalized chitosan3 h44.92Li et al. 2009
β-CD functionalized chitosan20.56
Cross-linked β-CD polymer41.11
Copper-based MOF HKUST-160 min (20°C)372Lin and Hsieh 2015
50 min (40°C)407
30 min (60°C)437
Pitch-based AC fibers24 h (25°C)313.95Liu et al. 2010
24 h (40°C)309.52
24 h (55°C)303.54
NH2-MIL-101(Al) MOF5 h192.67Liu et al. 2014a, 2014b
Fe3O4/mSiO2/GO120 min1548.78Liu et al. 2016a
Carbonized corncob residues30 min326.8 Liu et al. 2016b
ACs from BPL24 h330Lynam, Kilduff, and Weber 1995
Medium-rank Appalachian F400300
Grade bituminous coal FS100280
M15016 h (288 K)573.02Ma et al. 2014
16 h (303 K)519.46
16 h (318 K)424.39
Resin NDA-15016 h (288 K)636.88
16 h (303 K)604.07
16 h (318 K)572.87
XAD-416 h (288 K)582.22
16 h (303 K)524.39
16 h (318 K)529.95
ZrSiO4-MnO2 nanoparticles10 min12.09Mahmoud and Nabil 2017
Commercial AC PICA S23<3 days570.35Mohamed et al. 2011
Commercial AC F22431.24
Sludge-based AC347.78
AP-2.510 days127.3Moreno-Castilla et al. 1995
Commercial bentonite5 h0.37 Moussout et al. 2014
Chitosan flakes40 min0.63Ngah and Fatinathan 2006
Glutaraldehyde cross-linked chitosan beads2.48
Granular ACs: BDH4 days242Nouri 2002
Granular ACs: F100224
Granular ACs: SEI200
F1004 days (pH 2)222.58Nouri and Hagheseresht 2004
4 days (pH 12)66.77
H2-treated F1004 days (pH 2)225.36
4 days (pH 12)82.07
Urea-treated F1004 days (pH 2)214.23
4 days (pH 12)70.95
H2SO4-treated F1004 days (pH 2)186.41
4 days (pH 12)51.47
HDTMA-bentonite120 minNwokem et al. 2014
Briquette char7 days90Othman, Roddick, and Hobday 2000
Activated briquette char325
Power station char100
Activated power station char420
Activated Loy Yang brown coal370
Picactif from coconut shell550
Resin NDA-10024 h141.06Pan et al. 2005
Resin NA-01A24 hPan et al. 2006
TDTMA intercalated montmorillonite 6 hPark, Ayoko, and Frost 2011
Surfactant intercalated montmorillonite 12 hPark et al. 2013a
0.5 CEC HDTMA-montmorillonite12 h (23°C)46.08Park et al. 2013b
12 h (35°C)27.47
1.0 CEC HDTMA-montmorillonite12 h (23°C)33.67
12 h (35°C)23.20
1.5 CEC HDTMA-montmorillonite12 h (23°C)22.12
12 h (35°C)34.25
2.0 CEC HDTMA-montmorillonite12 h (23°C)25.45
12 h (35°C)40.32
Ag nanostructuresPerry et al. 2010
Synthesized nanozeolite150 min156.7Pham, Lee, and Kim 2016
Darco ACSabio et al. 2006
Coal fly ash390 min56.2 Sarkar and Acharya 2006
DP160 min (23°C)24.43Sarkar et al. 2010
60 min (37°C)26.57
DP260 min (23°C)42.15
60 min (37°C)42.43
CP160 min (23°C)36.45
60 min (37°C)29.35
CP260 min (23°C)36.45
60 min (37°C)30.47
AC24 hSerrano, Beltrán, and Segovia 1992
Microfibrous entrapped AC24 hShao, Zhang, and Yan 2013
Granular AC
Fly ash120 min7.8–9.6Singh and Nayak 2004
Clinoptilolite-type natural zeolite3 hSismanoglu and Pura 2001
Zn-Al/SDBS/SS-LDH24 h101.6Sun et al. 2014
AC fiber A1224 h (293 K)384.62Tang et al. 2007
24 h (308 K)370.37
24 h (323 K)357.14
Commercial unmodified and modified ACs7 daysTerzyk et al. 2008
Bismuth ferrocyanide6 h16.50Tewari 2014
Lead ferrocyanide21.0
Manganese ferrocyanide28.50
Zeolite240 min12.70Varank et al. 2012
Fe/Zn biochar30 h170Wang et al. 2017
Silica3 hWoods and Walker 2013
Aluminum MOF/RGO composite307.38Wu et al. 2016a
Cationic β-CD modified zeolite60 min0.25648Xiaohong et al. 2011
BDP-montmorillonite30 min76.92Xue et al. 2013
BDHP-montmorillonite40 min81.30
R. oryzae dead biomass40 min12.24Yaneva, Koumanova, and Allen 2013
Oxidized SWCNT30 min206Yao et al. 2014
Fe organo-inorgano-pillared montmorillonite24 hZermane et al. 2010
AC fiber24 h440.92Zhan et al. 2016
Polymer I60 min (303 K)8.53Zhang et al. 2014b
60 min (313 K)7.59
60 min (323 K)6.25
60 min (333 K)5.04
Nanographite oxide2 h268.5Zhang et al. 2015
3,5-Dinitrosalicylic acid-modified TiO210 minZheng et al. 2006
Chorella sp. Cha-0148 h82.2Zheng et al. 2017
Chlamydomonas sp. Tai-03110.7
Coelastrum sp. Pte-1580.1
Biochar from Chorella sp. Cha-01203.5
Biochar from Chlamydomonas sp. Tai-03105.9
Biochar from Coelastrum sp. Pte-15109.9
Powdered AC149.0
Montmorillonite6 h83.49Zhou et al. 2008
0.5 CEC HDTMA-pillared montmorillonite174.34
0.7 CEC HDTMA-pillared montmorillonite184.31
1.5 CEC HDTMA-pillared montmorillonite375.55
2.5 CEC HDTMA-pillared montmorillonite435.79
Fe2O3 synthesized with 0 g glycine24 h55.8Zhou et al. 2012
Fe2O3 synthesized with 1 g glycine79.2
Fe2O3 synthesized with 2 g glycine101.5
Fe2O3 synthesized with 2 g glycine74.5
Al2O3 synthesized with 0 mmol Na3C6H5O724 h137.0Zhou et al. 2013
Al2O3 synthesized with 0.125 mmol Na3C6H5O7185.2
Al2O3 synthesized with 0.25 mmol Na3C6H5O7200.0
Al2O3 synthesized with 0.5 mmol Na3C6H5O7181.8
Al2O3 synthesized with 1 mmol Na3C6H5O7217.4

  1. AP-2.5, AP-5, and AP-10, ACs from original coal; CEC, cation exchange capacity; Cloisite-10A, dimethylbenzyl-hydrogenated-tallow-modified SWy-1 montmorillonite; Cloisite-15A, dimethyl-dehydrogenated-tallow-modified SWy-1 montmorillonite; Cloisite-30B, methyl-bis-2-hydroxyethyl-tallow-modified SWy-1 montmorillonite modified; Cloisite-93A, methyl-dehydrogenated-tallow-modified SWy-1 montmorillonite; CP-5 and CP-10, ACs from demineralized coal; CP1 and CP2, 100% and 200% CEC equivalent cetylpyridinium chloride-palygorskite; CPC, cetylpyridinium chloride monohydrate; CS 1501 and RS 1301, commercial AC clothes; DP1 and DP2, 100% and 200% CEC equivalent dimethyldioctadecylammonium bromide-palygorskite; F100, commercial coal-based granular ACs; F400 and AQ40, commercial coal-based granular ACs; F400/100 and AQ40/100, commercial coal-based granular ACs ozonated at 100°C; F400/25 and AQ40/25, commercial coal-based granular ACs ozonated at 25°C; M150, magnetic hyper-cross-linked resin from copolymerization and post-cross-linking of tetraethoxysilane and vinyltriethoxysilane; NA-01A, dimethylamine-modified chloromethylated styrene-divinylbenzene cross-linked copolymer; NC 60, commercial granular AC; Polymer I, 4,4′-methylene-bis-phenyldiisocyanate cross-linked starch polymer; Qmax, maximum adsorption capacity; Resin NDA-150, commercial nonmagnetic resin; Spectracarb 2225, commercial AC cloth; TDTMA, tetradecyltrimethylammonium bromide; XAD-4, commercial nonmagnetic resin; Zn-Al/SDBS/SS-LDH, Zn-Al-layered double hydroxide modified with sodium dodecyl benzene sulfonate as surfactant and sodium salicylate as structure-directing agent.

Onto lignocellulosic materials and derivatives

Rhizopus oryzae dead biomass was used as an alternative environment-friendly adsorbent for p-NP removal from aqueous phase. This material showed 94% p-NP uptake efficiency, with the adsorption mainly limited by chemisorption and a nonnegligible role of intraparticle diffusion (Yaneva, Koumanova, and Allen 2013). On their part, Ahmaruzzaman and Gayatri (2011) investigated the adsorption of p-NP on activated neem leaves from simulated wastewater in batch and fixed-bed modes. The adsorption efficiency was higher in fixed-bed mode and adsorption best was described by the Freundlich model. The same authors equally showed that activated tea waste was a good adsorbent for the removal of p-NP. Adsorption of the latter onto activated tea waste best followed the Langmuir adsorption model and a pseudo-second order kinetics (Ahmaruzzaman and Gayatri 2010b). Rice husks (Ahmaruzzaman and Sharma 2005) and palm-tree fruit stones (Ahmed and Theydan 2012) have also been used as adsorbents for the uptake of p-NP from wastewaters.

Onto clay-like and modified clay structures

Pham, Lee, and Kim (2016) synthesized a nanozeolite adsorbent for the removal of p-NP. At the surface of this adsorbent, adsorption equilibrium was reached within 150 min with a maximum adsorption capacity of 156.7 mg/g. According to the authors, p-NP removal with the nanozeolite adsorbent was 46.6% cheaper than using AC. Nitric acid-activated kaolinite was prepared and used for the adsorption of p-NP with a maximum adsorption capacity of 0.281 mg/g (Azeez and Adekola 2016). Moussout et al. (2014) showed that p-NP is adsorbed onto commercial yellow bentonite with an adsorption capacity of 0.37 mg/g and equilibrium reached after 5 h. Other clay or clay-like samples have been used for the uptake of p-NP. These include cationic β-cyclodextrin (β-CD)-modified zeolite (Xiaohong et al. 2011), ZSM-11 zeolite (Lu et al. 2016), Fe-nanozeolite (Huong et al. 2016), clinoptilolite (Sismanoglu and Pura 2001), 1,3-bis(dodecyldimethylammonio)-propane dibromide (BDP)-montmorillonite and 1,3-bis(dodecyldimethylammonio)-2-hydroxypropane dichloride (BDHP)-montmorillonite (Xue et al. 2013), montmorillonite (Houari et al. 2014), hexadecyltrimethylammonium (HDTMA)-pillared montmorillonite (Zhou et al. 2008), Na-Y zeolite and kaolin clay (Ahmedzeki et al. 2013), zeolite and bentonite (Varank et al. 2012), HDTMA-bentonite (Koyuncu et al., 2011; Nwokem et al., 2014), dodecyltrimethylammonium-, benzyldimethyltetradecylammonium-, cetyltrimethylammonium-, octodecyltrimethylammonium-, and tetramethylammonium (TMA)-bentonite (Zhu, Chen, and Shen 2000), modified ZnAl-layered double hydroxide (Sun et al. 2014), MgAl-mixed oxide (Chen et al. 2009), poly(ethylene glycol) butyl ether (PEG)-bentonite (Koyuncu et al. 2011), and organopalygorskites (Sarkar et al. 2010) such as HDTMA-palygorskite (Chang et al. 2009).

Onto synthetic nanostructures

Surface-enhanced Raman spectroscopy and surface-enhanced infrared absorption were used to study the adsorption of p-NP onto host metal nanoparticles. p-NP was adsorbed onto nanoscale Ag film and powder as p-nitrophenolate ions, forming multilayers that depended on solvent polarity (Perry et al. 2010). Yao et al. (2014) explore the possibility of using pristine and modified carbon nanotubes (CNTs) as efficient adsorbents for the removal of p-NP from contaminated water. In their investigations, oxidized single-walled CNTs (SWCNTs) removed p-NP up to about 98%.

Graphene (Ismail 2015) and nanographite have been reported as efficient p-NP adsorbents. In this light, Liu et al. (2016a) synthesized a magnetic porous silica-GO hybrid composite (Fe3O4/mSiO2/GO) by grafting GO sheets onto the core-shell Fe3O4/mSiO2 nanoparticles and used for the removal of p-NP from aqueous solution. The adsorption process well fitted the pseudo-second-order kinetic model and the Langmuir isotherm model, with a maximum adsorption capacity of 1548.78 mg/g. Nanographite oxide prepared by the chemical oxidation method was also reported as an efficient adsorbent for the removal of p-NP by Zhang et al. (2015). The maximum adsorption capacity of nanographite oxide for p-NP was 268.5 mg/g at 283 K and pH 4.0.

Onto synthetic polymers

Ardelean, Davidescu, and Popa (2010) prepared a chloromethylated styrenedivinylbenzene copolymer with benzaldehyde groups for the removal of p-NP from aqueous solution. Amberlite XAD-4 was selected as comparative resin. The same group equally studied the uptake of p-NP at olefin-grafted poly(styrene-co-divinylbenzene) (Ardelean et al. 2012). Woods and Walker (2013) showed that the interactions between adsorbed p-NP and polymerized silica interface are dominated almost exclusively by close association with the silica substrate, with the aqueous solvent playing primarily a spectator’s role. Equilibration of adsorbed p-NP at the silica/aqueous interface required up to 3 h regardless of the pH of the aqueous phase. Li et al. (2009) studied the uptake of p-NP onto chitosan, salicylaldehyde functionalized chitosan, β-CD functionalized chitosan, and β-CD epichlorohydrin copolymer with varying adsorption efficiencies. These were 1.98, 44.92, 20.56, and 41.11 mg/g, respectively. Even better results were obtained by Ma et al. (2014) for the uptake of p-NP onto M150, NDA-150, and XAD-4 resins (Table 3). Other polymers reported for the uptake of p-NP include poly[2-(methacryloyloxy)ethyl]dimethylhexadecylammonium bromide (p-DMAC16) and poly(2-dimethylaminoethyl methacrylate) (p-DMA; Erdem et al. 2009), chitosan (Ngah and Fatinathan 2006), and various resins (Pan et al., 2006; Huang, Yan & Huang, 2009; Ma et al., 2014). Their performances are summarized in Table 3.

Onto metal-organic frameworks (MOFs)

MOFs have extra-high specific surface areas, highly ordered pore structures, and easy tunability of pore size and shapes (Furukawa et al. 2013). The aqueous phase adsorption of p-NP onto three MOFs, namely MIL-100(Fe), MIL-100(Cr), NH2-MIL-101(Al), and a reference AC at 303 K was studied by Liu et al. (2014a, 2014b). The adsorption capacities of MIL-100(Fe) and MIL-100(Cr) were limited, whereas NH2-MIL-101(Al) displayed an exceptional adsorption capacity toward p-NP (Table 3) through hydrogen bonding, exceeding that of the reference AC. Wu et al. (2016a) incorporated RGO into aluminum MOF [MIL-68(Al)]. This significantly changed the morphology of the MOF as well as increased its surface area. The new composite material showed high efficiency for the uptake of p-NP. The adsorption capacities reported were 175.44 mg/g (RGO), 271.00 mg/g [MIL-68(Al)], and 332.23 mg/g [MIL-68(Al)/RGO]. Recently, Chen et al. (2017) reported the removal of p-NP using the MOF Cr-BDC, with a rather low adsorption capacity of 19 mg/g. A much higher adsorption capacity was reported by Jia, Jiang, and Wu (2017) for the adsorption of p-NP onto amino-MIL-53(Al) (297.85 mg/g), whereas Lin and Hsieh (2015) reported even higher adsorption capacities (~400 mg/g) for the removal of p-NP using HKUST-1 MOF at different temperatures (Table 3).

Onto fibrous peat

The removal of p-NP from aqueous solution using a Brazilian fibrous peat was investigated by Jaerger et al. (2015). The adsorption capacity of the fibrous peat was shown to decrease with increasing temperature, with the adsorption process being predominantly chemisorption, exothermic and spontaneous.

Catalytic reductive degradation

Tang et al. (2015b) succeeded in degrading p-NP to p-aminophenol using nanoscale zero-valent Fe particles immobilized on mesoporous silica without ring opening. More than 80% p-NP was degraded by the synthesized catalyst after 30 days of exposure to air. In a similar work, Nakatsuji, Salehi, and Kawase (2015) completely degraded p-NP using zero-valent Fe within 30 min. Ma et al. (2016) proposed, in the same vein, a cobalt-appended 2D layered double oxide nanodisks for the reduction of p-NP.

p-NP degradation to p-aminophenol has also been investigated on RGO-cobalt oxide nanoparticles (RGO-Co3O4) nanocomposite (Al Nafiey et al. 2017), Trachyspermum ammi-supported biogenic Ag nanoparticles (Chouhan, Ameta, and Meena 2017), monometallic and bimetallic nanoparticles (Pozun et al. 2013), Fe/Ag bimetallic nanoparticles incorporated to p-aminothiophenol functionalized calcium alginate beads (Gupta et al. 2014), Ag/halloysite nanotubes/Fe3O4 nanocatalyst (Gan et al. 2015), pyrolusite (Peng et al. 2010), Ag nanoparticles loaded onto cellulosic fiber (Torkamani and Azizian 2016), hierarchical CNT membrane-supported Au nanoparticles (Wang, Dong, and Na 2013), SnO2 nanoparticles (Bhattacharjee and Ahmaruzzaman 2015a), 2D CuO nanoleaves (Bhattacharjee and Ahmaruzzaman 2015b), luminescent Au nanoclusters (Sinha and Ahmaruzzaman 2016), and superhydrophilic poly(vinylidene fluoride) membrane decorated with Au nanoparticles (Wu et al. 2016b).

Catalytic oxidative degradation

The almost complete removal of p-NP was achieved when Fe3+ supported on resin was used for its oxidative degradation in the presence of H2O2, as oxidant, in batch mode (Liou et al. 2010). Longli et al. (2005) showed that p-NP adsorbed onto granular AC was degraded during microwave irradiation. More than 90% p-NP was removed, with more than 65% mineralization. In another study, Qiu et al. (2016) employed bismuth oxide (Bi2O3) as catalyst in the microwave catalytic oxidation degradation of p-NP in polluted water. The total organic carbon removal ratio reached 99.74%. In the microwave catalytic oxidation degradation process, Bi2O3 was excited to generate electron-hole pairs by microwave irradiation, and the holes transformed H2O molecules adsorbed on the surface of the catalyst to hydroxyl radicals. The generated hydroxyl radicals then attacked p-NP molecules, converting them to CO2 and H2O. On their part, Xiong et al. (2016) degraded p-NP in aqueous solution using microsized Fe0/O3. Their p-NP degradation mechanism and pathway showed complete mineralization to CO2, H2O, and NO3 through a combination of homogeneous or heterogeneous catalytic ozonation, Fenton-like reaction, adsorption, and precipitation.

Photocatalytic methods

Complete photocatalytic degradation of p-NP over BiVO4 in the presence of H2O2, under visible light, was achieved by Umabala (2015) after 120 min. In the same vein, under UV light, p-NP was mineralized in the presence of TiO2 catalyst (Chen & Ray, 1998; Di Paola et al., 2003; Gota & Suresh, 2014; Islam et al., 2014) or Cr-Nb-TiO2 and Fe-Nb-TiO2 catalysts (Sikdar, Pathak, and Ghorai 2015). On their part, Hernández-Gordillo et al. (2014) degraded p-NP to p-aminophenol in the presence of hydrazine, under UV light irradiation, using either TiO2 or Ag-TiO2 as catalysts. In another study, the TiO2 catalyst was first chemically adsorbed onto 5-sulfosalicyclic acid before being employed for the photocatalytic degradation of p-NP (Li et al. 2005).

In the same light, Yang et al. (2010) degraded p-NP under artificial solar light using a Cu2O/TiO2 p-n junction catalyst. The photocatalytic activity of the latter was attributed to the extended absorption of visible light by the Cu2O nanowire network and the effective separation of photogenerated carriers driven by the photoinduced potential difference generated at the Cu2O/TiO2 p-n junction interface.

Fenton, photo-Fenton, and electro-Fenton degradation

The Fenton degradations employ the “Fenton reagent”, which is a combination of a ferrous salt and H2O2 (Lipczynska-Kochany 1992). They are advanced oxidation processes involving the hydroxyl radical as the oxidation agent (Zhang et al. 2007).

Fenton degradation of p-NP involves the following steps (Kavitha and Palanivelu 2005):


The generated hydroxyl radicals then degrade p-NP to mineralized products:


In photo-Fenton degradation, there is increased efficiency of the process as hydroxylated ferric ions are photoreduced to continuously supply hydroxyl radicals (Kavitha and Palanivelu 2005):


Using a comparative approach, Kavitha and Palanivelu (2005) degraded p-NP and three other nitrophenols using the Fenton, solar-Fenton, and UV-assisted Fenton processes. With the Fenton process, only about 32% degradation was reported, whereas with the solar-Fenton and UV-assisted processes more than 92% mineralization was obtained. In other comparative studies, it was equally observed that photo-Fenton treatment led to more rapid degradation of p-NP than Fenton treatment (Goi & Trapido, 2002; Kiwi, Pulgarin & Peringer, 1994).

On its part, electro-Fenton degradation involves the formation of H2O2 through a two-electron cathodic reduction of dissolved oxygen (Ammar, Oturan, and Oturan 2007). The process can be summarized as follows:

O2+2H++2eH2O2 (at the cathode)
Fe2++H2O2+H+Fe3++H2O+OH (in solution)
Fe3++eFe2+ (at the cathode)

In this light, Oturan et al. (2000) electrocatalytically generated Fenton’s reagent for the production of hydroxyl radicals at a carbon felt working electrode held at −0.5 V versus a saturated calomel reference electrode. In the presence of aqueous p-NP, the authors obtained complete mineralization of the latter as well as its intermediate degradation products. This complete mineralization was ascertained through total organic carbon analyses. Complete mineralization of p-NP was also obtained by Zhang et al. (2007) using electro-Fenton degradation in batch recirculation mode.


The use of microorganisms for pollutant remediation is advantageous because no harmful intermediates are generated, there is complete pollutant destruction, and the process is environment friendly and economical (Timmis, Steffan & Unterman, 1994; Soccol et al., 2003; Kulkarni & Chaudhari, 2007). The remediation of p-NP by various microorganisms in both aerobic and anaerobic conditions is summarized in Table 4 and Table 5.

Table 4:

Aerobic biodegradation of p-NP.

MicroorganismDegradation pathwayReference
Bacillus licheniformis CFR 1021, Xanthomonas maltophila CFR 1022, Serratia liquefaciens CFR 1023, Pseudomonas putida CFR 1711, Pseudomonas sp. CFR 1712, Pseudomonas alcaligenes CFR 1713, Pseudomonas sp. CFR 1714. and Sarcina maxima MTCC 5216 ConsortiumFormation of 4-nitrocatechol, 1,2,4-benzenetriol, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipateBasheer et al. 2007
S. maximaFormation of HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipate
P. putidaFormation of nitriteBhatti, Toda, and Furukawa 2002
Arthrobacter protophormiae, Burkholderia cepaciaRelease of nitrite, formation of HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipate, TCA cycleBhushan et al. 2000
Ralstonia sp.Release of nitrite, formation of 1,2,4-benzenetriol
protophormiaeRelease of nitrite, formation of HQ, γ-hydroxymuconic semialdehyde, β-ketoadipate, TCA cycleChauhan, Chakraborti, and Jain 2000
Burkholderia sp.Formation of p-nitrocatechol, 1,2,4-benzenetriol, maleylacetate, and β-ketoadipate, TCA cycleChauhan et al. 2010
Nocardioides NSP41Simultaneous degradation of p-NP and phenolCho, Rhee, and Lee 2000
Pseudomonas aeruginosa and Bacillus cereus from wastewater microfloraDaffri, Harzellah, and Bousseboua 2014
Moraxella G21Complete mineralizationErrampalli et al. 1999
Rhodococcus wratislaviensisRelease of nitrite, formation of HQ, γ-hydroxymuconic semialdehyde, and β-ketoadipate, TCA cycleGemini et al. 2005
Pseudomonas fluorescensRelease of nitrite, mineralization to CO2Heitkamp et al. 1990
ActinomyceteRelease of nitrite, mineralization to CO2Herman and Costerton 1993
Arthrobacter sp.Release of nitrite, formation of 4-nitrocatechol or 4-nitroresorcinol, 1,2,4-benzenetriol, maleylacetate, and β-ketoadipateJain, Dreisbach, and Spain 1994
Bacillus sphaericus JS905Release of nitrite, formation of 4-nitrocatechol and 1,2,4-benzenetriol, maleylacetate to 3-ketoadipic acidKadiyala and Spain 1998
Rhodococcus opacusFormation of HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipate, TCA cycleKitagawa, Kimura, and Kamagata 2004
Formation of p-nitrocatechol, hydroxyquinol, maleylacetate, and β-ketoadipate, TCA cycle
P. putidaComplete degradation with release of nitriteKulkarni and Chaudhari 2006
protophormiaeComplete degradation in soilLabana et al. 2005a
protophormiaeComplete degradation in soil microcosmLabana et al. 2005b
Sphingomonas sp.Release of nitrite, mineralization to CO2Leung et al. 1997
Pseudomonas WBC-3Complete degradation of p-NPLieu et al. 2005
Stenotrophomonas sp.Formation of HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipate, TCA cycleLiu, Yang, and Qiao 2007
Ralstonia eutrophaFormation of p-benzoquinone, HQ to aerobic respirationMaleki et al. 2015
Formation of p-nitrocatechol, p-benzoquinone, and HQ to aerobic respiration
Pseudomonas sp.Release of O2 and NO2, formation of benzoquinone, HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipate, TCA cycleMattozzi and Keasling 2010
Serratia sp.Formation of p-nitrocatechol, 1,2,4-benzenetriol, and maleylacetate, TCA cyclePakala et al. 2007
Ochrobactrum sp.Release of nitriteQiu et al. 2007
Pseudomonas sp. and bacterial consortiumFormation of HQ and release of nitriteQureshi and Purohit 2002
P. putidaRelease of nitriteRay, Ait, and Loser 1999
P. putidaFormation of HQSamuel, Sivaramakrishna, and Mehta 2014
PseudomonasMineralization to CO2Schmidt, Scow, and Alexander 1987
Bacillus sp.Formation of p-nitrocatechol to HQ to TCA cycleSengupta, Maiti, and Saha 2015
Moraxella sp.Release of nitrite, formation of HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipateSpain and Gibson 1991
Moraxella sp.Release of nitrite, mineralization to CO2Spain, Wyss, and Gibson 1979
Phanerochaete chrysosporiumDegraded to 1,2-dimethoxy-4-nitrobenzene via 4-nitroanisoleTeramoto, Tanaka, and Wariishi 2004
Corynebacterium sp., Pseudomonas sp.MineralizationZaidi and Mehta 1995
Pseudomonas sp.Formation of benzoquinone, HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipate, TCA cycleZhang et al. 2009
Moraxella sp.Formation of benzoquinone, HQ, γ-hydroxymuconic semialdehyde, maleylacetate, and β-ketoadipate, TCA cycleZhang et al. 2012b
R. opacusFormation of p-nitrocatechol, hydroxyquinol, maleylacetate, and β-ketoadipate, TCA cycle

  1. HQ, hydroquinone; TCA, tricaboxylic acid.

Table 5:

Anaerobic biodegradation of p-NP.

MicroorganismDegradation pathwayReference
Flocs/sludge granules of methanogens4-Nitrophenol converted to corresponding amineDonlon et al. 1996
Methanobacterium formicium, Desulphatomaculem orientisComplete transformationGorontzy, Kuver, and Blotevogel 1993
Methanogenic bacteriaComplete degradationHaghighi-Podeh and Bhattacharya 1996
Haloanaerobicum praevalens, Sporohalobacter marismortuiReduction to aminophenolOren, Gurevich, and Henis 1991
Granular sludge from UASB reactorAccumulation of the corresponding aminophenol as dead-end productRazo-Flores et al. 1997
Granulated anaerobic sludge from UASB reactorMineralized to methaneSponza and Kuscu 2005

  1. UASB, upflow anaerobic sludge blanket.


Kotronarou, Mills, and Hoffmann (1991) studied the sonolytic degradation of p-NP in aqueous solution using a frequency of 20 kHz. The degradation products found were NO2, NO3, and H+ as primary products and HCOO, C2O42−, and H2O2 as secondary products. They also observed a linear increase in the concentration of H2O2 during the sonolysis process. The degradation path involved both hydroxyl radical (OH) attack and pyrolytic reactions. With 16 and 20 kHz frequency magnetostrictive transducers, Hua, Hochemer, and Hoffmann (1995) degraded p-NP on a larger scale in a continuous flow operation (3.2 l/min flow rates). In another study, Sivakumar, Tatake, and Pandit (2002) sonochemically degraded p-NP in aqueous solution using three operating frequencies (25, 40 and 25 + 40 kHz) and found an improvement in the rate of degradation in the combined frequency mode (25 + 40 kHz).

Xu, Shi, and Wang (2005) combined sonolysis and ozonolysis to obtain enhanced p-NP degradation rates, higher than those obtained with ultrasound or ozone individually. In this approach, ozone is decomposed by OH or HO2 ions to OH and HO2 radicals, whereas H2O, in the sonolysis system, is pyrolytically decomposed to OH and HO2 radicals during acoustic cavitation. The free radicals produced and ozone then attack p-NP, leading to its degradation directly inside the cavitation bubble or in its interfacial sheath. Earlier, Weavers, Ling, and Hoffmann (1998) had proposed a conceptual pathway showing the interactions between ozonolysis and sonolysis during the degradation of p-NP (Figure 3).

Figure 3: Conceptual diagram of possible pathways of p-NP degradation and interactions of sonolysis and ozonolysis (adapted from Weavers, Ling, and Hoffmann 1998).© Copyright American Chemical Society and reproduced with permission.

Figure 3:

Conceptual diagram of possible pathways of p-NP degradation and interactions of sonolysis and ozonolysis (adapted from Weavers, Ling, and Hoffmann 1998).

© Copyright American Chemical Society and reproduced with permission.

Other remediation methods

Wu et al. (2005) reported a combination of electrocatalysis and AC in fluidization mode for the abatement of p-NP. With such a synergetic approach, they were able to remove 150 mg/l p-NP in no more than 30 min due to the formation of AC microelectrodes under the electric field. AC played the double role of adsorbent and catalyst. A similar study was reported by Zhou and Lei (2006). The advantages of such a combined approach are enhanced cost-effectiveness of the entire process by reducing equipment and flow and regeneration of AC to minimize fresh supplies.

On their part, Follut, Karpel, and Leitner (2007) degraded p-NP by radiolysis. By this approach, aqueous solutions of p-NP containing TiO2 or Al2O3 were irradiated by electron beams. This electron beam irradiation process led to the formation of 4-nitrocatechol, nitrite, and nitrate ions as degradation products.

Electrochemical degradation of p-NP at boron-doped diamond and platinum anodes was studied by Zhang et al. (2013b). Such degradation was possible thanks to an in situ concomitant generation of hydroxyl radicals and chlorine-based oxidant species. The degradation products (HPLC determined) and pathways proposed by the authors are shown in Figure 4.

Figure 4: Degradation pathways of p-NP at (A) boron-doped diamond and (B) platinum electrodes (Zhang et al. 2013b; © Copyright Elsevier BV).Reproduced with permission.

Figure 4:

Degradation pathways of p-NP at (A) boron-doped diamond and (B) platinum electrodes (Zhang et al. 2013b; © Copyright Elsevier BV).

Reproduced with permission.

Ali et al. (2006) developed a solid-phase extraction method for p-NP removal from wastewater. The elution was carried out by methanol at 0.25 ml/min flow rate and about 93% recovery registered.

Conclusion and outlook

In this review, we summarized the progress made in the last decades to detect, quantify, and eliminate p-NP from wastewaters and other contaminated ecosystems. A broad spectrum of analytical tools exists for the elaboration of p-NP sensors and adsorbents as well as abatement. Some of them still, however, pose a serious environmental problem.

The ideal remains the complete mineralization of p-NP. This should be coupled with the development of cheap natural products that can neutralize the acidic solutions obtained after such complete mineralization. In this regard, the use of techniques such as microwave irradiation, sonolysis, and bioremediation is to be encouraged. Concerning bioremediation, emphasis should be laid on developing microbial strains that completely mineralize p-NP. Such eco-friendly approaches shall certainly meet wide public acceptance. The future also lies in the combination of methods for cost-effectiveness and reduced equipment as well as the development of permeable reactive barriers for in situ on treatment zones.


Financial support from The World Academy of Sciences for the Advancement of Science in Developing Countries (TWAS grant no. 12-117 RG/CHE/AF/AC-G awarded to I.K. Tonle) is gratefully acknowledged. The authors also thank the International Science Programme (ISP, Sweden) for its support to the African Network of Electroanalytical Chemists (ANEC).


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Received: 2017-8-25
Accepted: 2018-1-30
Published Online: 2018-3-24

©2018 Walter de Gruyter GmbH, Berlin/Boston

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