Introduction

Chromium compounds are environmental contaminants present in ground water, soil and industrial effluents due to their extensive use in various industries (Madhavi et al. 2013). The trivalent Cr(III) and hexavalent Cr(VI) are the most stable and abundant forms in the environment (Cervantes et al. 2001); the common Cr(VI) species present in aqueous solution are hydrochromate (HCrO4 −), chromate (CrO4 2−) and dichromate (Cr2O7 2−), depending on the solution pH, the Cr(VI) concentration and on the redox potential. On the other hand, Cr(III) derivatives are much less mobile and exist in the environment mostly forming stable complexes with both organic and inorganic ligands; in aqueous solution at neutral pH, Cr(III) tend to associate with hydroxide OH− ions in a pH-dependent manner and precipitates as hydroxide [Cr(OH)3] or hydrated oxide (Cr2O·H2O) (Barrera-Díaz et al. 2012; Ramírez Díaz et al. 2008). The biological effects of Cr strongly depend on its oxidation state and cellular localization. Cr(VI) is considered the most toxic form of Cr for microorganisms and mammalian cells. In the intracellular medium Cr toxicity is mainly related to the process of reduction of Cr(VI) by the action of flavoenzymes or metabolites (i.e. ascorbic acid, glutathione) to produce the active intermediates Cr(V) and/or Cr(IV) and the reactive oxygen species (ROS) peroxide, hydrogen peroxide and hydroxyl radicals, with Cr(III) as the final product. ROS production causes oxidative damage to DNA, proteins and lipids and Cr(III) affects DNA replication, causes mutagenesis, and alters enzyme structure and activity by reacting with their carboxyl and thiol groups (Viti et al. 2014; Ramirez-Diaz et al. 2008). Bacteria of different genera (for example, Arthrobacter, Cupriavidus, Escherichia, Pseudomonas, Shewanella) and few cases of fungi (the yeast Saccharomyces cerevisiae) have been used in transcriptomic and/or proteomic studies on their responses to Cr(VI). Overall, the results obtained indicated changes in functions related to central metabolism and energy production and conversion, as well as effects on different transporters and DNA metabolism and repair (Viti et al. 2014).

Microbial interactions with chromium

Microbial mechanisms of the interaction with Cr are of both fundamental and biotechnological interest, because microorganisms have a variety of properties that can affect changes in Cr speciation, toxicity, and mobility. Therefore, these microbial mechanisms of Cr detoxification may serve as the basis for the development of new technologies to remove Cr from contaminated settings.

Some previous articles have reviewed the information published regarding the mechanisms of chromate resistance in microorganisms and/or those related to the enzymatic and non-enzymatic processes of Cr(VI) reduction (Cheung and Gu 2007; Ramírez-Diaz et al. 2008; Poljsak et al. 2010; Viti et al. 2014; Thatoi et al. 2014; Joutey et al. 2015). In this overview, our aim was to summarise the recent literature addressing the different microbial mechanisms of interactions with Cr, in an attempt to connect the basic biological processes to biotechnological applications.

Microbial mechanisms of the interactions with Cr involve active or passive processes, which are schematically shown in Fig. 1. These processes include transport, accumulation, biosorption, and the enzymatic and non-enzymatic reduction of the oxyanion to Cr(III).

Fig. 1
figure 1

Schematic summary of microbial interactions with chromium

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Chromium transport

Due to the structural similarity of Cr(VI) with SO4 2− anion, it can easily be transported across biological membranes via active sulphate transporters in bacteria and fungi (Cervantes et al. 2001) (Fig. 1). Another chromate transport system is represented by the ChrA protein, which belongs to the chromate ion transporter (CHR) superfamily that includes hundreds of homologues from all three domains of life. ChrA protein is composed of two families of sequences: the short-chain monodomain family and the long-chain bidomain family (Díaz-Pérez et al. 2007; Viti et al. 2014). The efflux of chromate is a resistance mechanism in bacteria conferred by the ChrA protein, which functions as a chemiosmotic pump that effluxes chromate from the cytoplasm using the proton motive force (Ramírez Díaz et al. 2008; Viti et al. 2014). ChrA homologues in fungi appear to function in an opposite manner to those in bacteria. In the filamentous fungus Neurospora crassa, the CHR-1 protein is a bidomain member of the CHR superfamily, which acts as a transporter of chromate favouring the accumulation of the oxyanion in cells, resulting in chromate sensitivity (Flores-Alvarez et al. 2012).

Besides the classic chromate transport systems such as sulphate (chromate) permeases, another chromate transport system was described in the yeast Saccharomyces cerevisiae, which is limited by actin-mediated endocytosis. This system was proposed to involve the ubiquitin-dependent endocytic inactivation of plasma membrane Cr transporter(s), independent of the sulphate (and chromate) permeases Sul1p and Sul2p (Holland and Avery 2009). Figure 1 summarises the action of these various chromate transport systems, highlighting the differences between bacteria and fungi.

By contrast, in microbial cells Cr(III) crosses cell membranes with low efficiency because it tends to form insoluble compounds. However, it has also been observed that microbial reduction of Cr(VI) in the extracellular medium may transiently generate soluble Cr(III) in the form of Cr3+ ion and/or hydroxyl complexes which can have deleterious effects on cells (Bencheikh-latmani et al. 2007). In any case, the transporter responsible for incorporating Cr(III) into microbial cells has not been identified to date, although this ion can be formed inside cells by either enzymatic or non-enzymatic Cr(VI) reduction (Fig. 1).

Bioaccumulation of chromium

Microorganisms are capable of incorporating metals into their cells through two processes, one called “passive uptake” or “biosorption” and the other known as “active uptake”. The former process is considered as the metabolism-independent accumulation of metals by living or inactive non-living biomass or biological materials, whereas active uptake, which occurs only in living cells, requires metabolism and energy for the transport of metals across the cell membrane into the cells. The combination of the active and passive modes of metal uptake is called “bioaccumulation” (Mohan and Pittman 2006; Wang and Chen 2009).

The cell wall of microorganisms is the first component that comes into contact with metal ions, and is therefore largely responsible for the biosorption of metals (Wang and Chen 2009). The cell wall chemical functional groups participating in Cr binding have been identified using different analytical techniques, including Fourier transform infrared spectroscopy (FTIR), scanning/transmission electron microscopy with energy dispersive X-ray spectroscopy (SEM-EDX/TEM-EDX), X-ray diffraction (XRD) analysis, X-ray absorption spectroscopy (XAS) and X-ray photoelectron spectroscopy (XPS) (Mukherjee et al. 2013) (Table 1). For instance, the use of FTIR revealed the functional groups involved in Cr(VI) binding in the cell walls of the bacterium Streptomyces werraensis LD22, including amine, hydroxyl, and carboxyl. The same technique indicated that the functional groups involved in Cr(VI) binding in the cell walls of the fungus Aspergillus niger were hydroxyl, carboxyl, amino, and carbonyl (Table 1).

Table 1 Functional chemical groups related to chromium binding with microbial cell walls and analytical techniques employed
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The adsorption coupled reduction of Cr(VI) is a process by which Cr(VI) is reduced to Cr(III) in the aqueous phase or in the biomass via contact with electron-donor groups of the biomass under an acidic pH (Park et al. 2006; Fig. 1); this process has been demonstrated in fungi (Park et al. 2007) and bacteria (Polti et al. 2011) using analytical techniques such as XPS, XAS, and SEM-EDX/TEM-EDX. However, several studies have also demonstrated that some bacteria species such as cyanobacteria (Ozturk and Aslim 2008), Azotobacter (Joshi and Juwarkar 2009), Arthrobacter (Shuhong et al. 2014), and Bacillus (Dogan et al. 2015), as well as fungal species such as Trichoderma (Chang et al. 2016) and Schwanniomyces (Mohite et al. 2015) produced exopolymeric substances (EPS) with the capacity to remove Cr(VI) by an adsorption coupled reduction process.

Recently, it has been reported that chromium oxide nanoparticle synthesis follows Cr(VI) reduction in the Bacillus cereus strain XMCr-6 (Dong et al. 2013) and the biosorption process in Bacillus subtilis cells (Annamalai et al. 2014) exposed to Cr(VI); in both studies, the size, morphology, and location of the biosynthesised nanoparticles were determined by SEM, TEM, XRD, or FTIR. In the yeast Schwanniomyces, the presence of EPS generates nanoparticles that show a coupled adsorption-reduction system (Mohite et al. 2015). Table 2 summarises the results obtained in the last 9 years on Cr removal by biosorption with bacteria and fungi.

Table 2 Cr(VI) removal by biosorption with bacterial and fungal biomass
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Enzymatic reduction

In bacteria, the enzymatic Cr(VI) reduction system can be attributed to the catalysis performed either by soluble cytosolic proteins or insoluble cell membrane enzymes (Ramirez-Diaz et al. 2008; Thatoi et al. 2014; Viti et al. 2014) (Fig. 1). Numerous bacterial genera, including Pseudomonas, Bacillus, and Arthrobacter, have been widely reported to reduce Cr(VI) using an enzymatic process, either aerobically or anaerobically or under both conditions (Thatoi et al. 2014; Viti et al. 2014). In Pseudomonas species, the genes related to Cr(VI) reduction are located in plasmids, whereas in several bacilli and Enterobacteriaceae species, these genes are located on the chromosomal DNA (Thatoi et al. 2014).

Some bacterial chromate reductases such as ChrR, YieF, NemA, and LpDH have been identified, which are located either in soluble fractions (cytoplasm) or bound to the membrane of the bacterial cell; the conditions under which these enzymes are functional can be aerobic or anaerobic, or sometimes both.

Chromate reductase activity has been studied in the fungi Penicillium sp. (Arévalo-Rangel et al. 2013) and A. niger (Gu et al. 2015). In both cases, it was found that the chromate reductase was mainly located in the soluble fraction of cells. In Aspergillus flavus, chromate reductase activity is accompanied by a biosorption process, indicating that there is a complex remediation mechanism of Cr(VI), which includes an interconnection of different interactions with the metal (Singh and Bishnoi 2015). Cr(VI) reduction can also be achieved indirectly by enzymes such as glucose oxidase from A. niger, which produces molecules such as gluconolactone and hydrogen peroxide that can reduce Cr(VI) to Cr(III) (Romo-Rodríguez et al. 2015) (Fig. 1).

In addition, some recent studies have focused on determining the ability of Cr(VI) reduction by chromate reductase enzymes present in the supernatant of microbial cultures. Rath et al. (2014) reported high extracellular chromate reductase activity under an optimised set of conditions from a Bacillus amyloliquefaciens strain isolated from chromite mine soil. An isolate of Arthrobacter sp. from this same environment showed extracellular chromate reductase activity to reduce the Cr(VI) present in the chromite mine effluent (Dey and Paul 2016). In B. cereus RMLAU1, which catalyses the co-remediation of pentachlorophenol and chromate, chromate reductase activity was observed in the culture supernatant, cytosolic fraction, as well as in cell debris under aerobic conditions; the maximum enzyme activity was found in the cytosolic fraction (48 %), followed by that in the culture supernatant (39.7 %) and cell debris (12.3 %) (Tripathi et al. 2013).

Non-enzymatic reduction

Reduction of Cr(VI) may take place by chemical reactions associated with compounds present in intra/extracellular compounds produced during microbial metabolism, including amino acids, nucleotides, sugars, vitamins, organic acids, or glutathione (Cervantes et al. 2001; Dhal et al. 2013) (Fig. 1). For instance, ascorbate is capable of reducing Cr(VI), and the riboflavin derivatives FAD and FMN are essential coenzymes for chromate-reducing flavoenzymes (Cervantes et al. 2001). In addition, it is well known that Fe(II) and H2S, the anaerobic metabolic end products of iron- and sulphate-reducing bacteria (SRB), can reduce hexavalent Cr, both individually and in combination (Somasundaram et al. 2009).

In bacteria, direct Cr(VI) reduction has been demonstrated by means of the production of the siderophore pyridine-2,6-bis (thiocarboxylic acid) (pdtc) by Pseudomonas stutzeri KC, which reduces 86 % of hexavalent Cr. In addition, the pdtc hydrolysis products pyridine-2-carboxylic-6-thiocarboxylic acid and dipicolinic acid also reduce Cr(VI) (Zawadzka et al. 2007).

Excreted fungal metabolites such as the organic acids citrate (Barrera-Diaz et al. 2012) and oxalate (Barrera-Diaz et al. 2012; Wrobel et al. 2015) have also been reported to participate in the reduction of Cr(VI) through the photocatalytic effect of Fe(III) or by Mn without light (Fig. 1). The importance of this process in a biological context was illustrated by the finding that the extracellular reduction of Cr(VI) to Cr(III) in the presence of Mn2+ is negatively affected in the oxalic acid-non producer mutants of Aspergillus tubingensis Ed8 (Fig. 2) (unpublished data).

Fig. 2
figure 2

Cr(VI) reduction in A. tubingensis Ed8 and the null mutants Δoah IT4-9 and IT5-13. The Cr(VI) reduction capacity of wild-type A. tubingensis Ed8 (black squares) and its derived null mutants in the oah gene, IT4-9 (dark grey squares) and IT5-13 (light grey squares), at 24 h of culture was determined by the diphenyl-carbazide method. The culture medium was supplemented with 30 mg L−1 Cr(VI) and 5 mM MnCl2

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Biotechnological applications

Of the different mechanisms of microbial interaction with chromium biosorption and bioreduction are those that have been considered for the development of biotechnological strategies for the removal of Cr from the environment. As shown in Table 3, several systems have been implemented for use in different types of reactors for Cr removal, using monocultures or biomass of bacterial or fungal strains, as well as consortia of different bacteria or fungi. The use of microbial consortia in some of the systems relies on the notion that consortia of microorganisms are metabolically superior for removing metals and are more suitable for field applications (Joutey et al. 2015). Table 3 shows the results of an ex situ study for Cr(VI) bioremediation using a native consortium of SRB carried out in an up-flow anaerobic sludge bed bioreactor, which showed a 90 % removal efficiency by reduction of 50 mg L−1 Cr(VI) (Qian et al. 2016). Table 3 also highlights the use of a consortium of bacteria (Acinetobacter junii, Escherichia coli, B. subtilis) incubated in a continuous packed bed column reactor, which improved the Cr(VI)-reducing capability as compared to that of a batch reactor. In several Cr-contaminated environmental water matrices with 100 mg L−1 Cr(VI), the consortium showed 55–65 % efficiency of Cr(VI) removal by adsorption (Samuel et al. 2013).

Table 3 Biotechnological approaches for the removal of Cr(VI)/Cr(III)
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Another biotechnological process of interest is based on the bacterial cells and enzymes immobilised in different polymer matrices such as agar layers, polyhydroxyalkanoate granules, calcium alginate, alginate beads, and poly(vinyl alcohol) alginate, which have proven to be effective for Cr(VI) reduction. Table 3 shows the performance of the ChromeBac™ system, based on the use of Acinetobacter haemolyticus EF369508, which was immobilised onto rubber wood sawdust inside a bioreactor; after 90 % of the initial Cr(VI) (30–60 mg L−1) was reduced, the residual Cr(VI) was reduced to between 1.0 and 1.5 mg L−1 by immobilised chromate reductase alginate beads packed in a flow-through column (Ishak et al. 2016).

Two types of bioreactors, a stirred tank bioreactor and a concentric tube airlift bioreactor, were compared for the removal efficiency of Cr by the filamentous fungus Trichoderma viride, and the results showed that the airlift bioreactor might be a promising alternative, especially when shear-sensitive microorganisms are used (Morales-Barrera and Cristiani-Urbina 2006) (Table 3). The ability of the yeast Candida tropicalis to remove Cr was tested in artificially contaminated soils in a microcosm system to simulate natural environmental conditions, showing reduction of 72.2 % of 40 mg L−1 Cr(VI) (Bahafid et al. 2013). An A. niger strain isolated from a tannery was used in an airlift reactor for the treatment of tanning wastewater, leading to maximum removal efficiency by adsorption of 88 % of an initial Cr(III) concentration of 1300 mg L−1 (Sepehr et al. 2012a). These results are even better than those obtained (removal efficiency of 72 %) with the same organism using a stirred tank reactor (Sepehr et al. 2012b). The use of a consortium of Cr-resistant fungi immobilised in a support material in a stirred tank bioreactor filled with wastewater from a tannery achieved an overall removal efficiency of 99.9 % of the total Cr present (Sharma and Malaviya 2016).

Other systems with potential applications in chromium removal

As mentioned above, the application of immobilised cells in Cr bioremediation has proven to be satisfactory in some cases. This process has also been favoured along with the development of nanotechnology. Pang et al. (2011) demonstrated that carbon nanotubes impregnated into Ca-alginate beads, which were used to immobilise Pseudomonas aeruginosa cells, markedly enhanced the stability of the enzyme and the process of Cr(VI) reduction. Another process related to nanotechnology is the mobilisation and immobilisation of metal ions by bacteria at the nanometer scale. In this regard, a process for remediating Cr-contaminated sites has been developed by applying metal-reducing bacteria in combination with nano-minerals with a size ranging from 10 to 50 nm [siderite (FeIICO3)], which were obtained from acid mine drainage and act as electron donors for the catalysis of Cr(VI) reduction and immobilisation (Cr(III)-containing precipitates) (Seo and Roh 2015).

In relation to cell-immobilisation processes, Robins et al. (2013) applied genetic engineering techniques to develop a system for Cr(VI) remediation based in an immobilised chromate reductase. They constructed a fusion of the E. coli nemA gene and the polyhydroxyalkanoate synthase gene phaC from Ralstonia eutropha, which enabled the high-level biosynthesis of functionalised polyhydroxyalkanoate granules displaying stable and active NemA chromate reductase on their surface. They proposed that this system offers good promise as an economic solution for ex situ Cr(VI) remediation. In the same sense, Barak et al. (2006) obtained an E. coli strain that produces a mutant enzyme termed ChrR6. This enzyme was obtained by a directed evolution approach through the error-prone polymerase chain reaction technique, and showed 200-fold greater chromate-reducing activity than the wild-type E. coli ChrR enzyme. Collectively, these studies demonstrate that enzymes from genetically modified bacteria represent a promising approach for the bioremediation of Cr(VI)-contaminated wastewater over a wide range of environmental conditions.

Concluding remarks

Bacteria and fungi display diverse mechanisms of interactions with Cr. The uptake of chromate through the sulphate transporter is a common process in bacteria and fungi; however, these organisms appear to differ with respect to the function of the chromate transporter homologous proteins ChrA/ChrR. Both bacteria and fungi show Cr(VI)-reducing capability, although the pathways of enzymatic and non-enzymatic Cr(VI) reduction have been predominantly studied in bacteria, and the location and identity of the enzymes and metabolic processes involved have been elucidated. Nevertheless, a few studies have described Cr(VI)-reducing activities and metabolic processes in fungi. Bacteria and fungi have the ability to accumulate Cr in the biomass, and the cell wall functional groups involved in Cr binding have been determined in several cases. Cr biosorption and bioreduction are the primary mechanisms that have attracted attention for the development of biotechnological processes for Cr removal. Several systems have implemented the use of different types of reactors for Cr removal using monocultures or consortia of bacteria or fungi. The application of immobilised microbial cells combined with the development of nanomaterials is facilitating development of processes for Cr remediation. In this context, studies have been performed with recombinant microorganisms overexpressing the chromate reductase and other enzymes.