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Fe(III) minerals play a crucial role for arsenic (As) mobility in aquifers as they usually represent the main As-bearing phases. Microbial reductive dissolution of As-bearing Fe(III) minerals is responsible for the release of As and the resulting groundwater contamination in many sites worldwide. So far, in most studies mainly abiogenic iron minerals have been considered. Yet, biogenic minerals that possess different properties to their abiogenic counterparts are also present in the environment. In some environments they dominate the iron mineral inventory but so far, it is unclear what this means for the As mobility. We, therefore, performed an in-situ aquifer Fe(III) minerals exposure experiment i) to evaluate how different biogenic and abiogenic Fe(III) minerals are transformed in a strongly reducing, As-contaminated aquifer (25 m) compared to As-free moderately reducing aquifer (32 m) and ii) to assess which microbial taxa are involved in these Fe(III) minerals transformations. We found that higher numbers of bacteria and archaea were associated with the minerals incubated in the As-contaminated compared to the non-contaminated aquifer and that all Fe(III) minerals were mainly colonized by Fe(III)-reducing bacteria, with Geobacter being the most abundant taxon. Additionally, fermenting microorganisms were abundant on minerals incubated in the As-contaminated aquifer, while methanotrophs were identified on the minerals incubated in the As-free moderately reducing aquifer, implying involvement of these microorganisms in Fe(III) reduction. We observed that biogenic Fe(III) minerals generally tend to become more reduced and when incubated in the As-contaminated aquifer sorbed more As than the abiogenic ones. Most of abiogenic and biogenic Fe(III) minerals were transformed into magnetite while biogenic more crystalline mixed phases were not subjected to visible transformation. This in-situ Fe(III) minerals incubation approach shows that biogenic minerals are more prone to be colonized by (Fe(III)-reducing) microorganisms and bind more As, although ultimately produce similar minerals during Fe(III) reduction.
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Microbial transformation of biogenic and abiogenic Fe minerals
followed by in-situ incubations in an As-contaminated vs.
non-contaminated aquifer
Martyna Glodowska
, Magnus Schneider
, Elisabeth Eiche
, Agnes Kontny
Thomas Neumann
, Daniel Straub
, AdvectAs project members
, Sara Kleindienst
Andreas Kappler
Geomicrobiology, Center for Applied Geosciences, University of Tübingen, Germany
Microbial Ecology, Center for Applied Geosciences, University of Tübingen, Germany
Department of Microbiology, IWWR, Radboud University, the Netherlands
Karlsruhe Institute of Technology, Institute of Applied Geosciences, KIT, Germany
Technical University of Berlin, Institute for Applied Geosciences, Berlin, Germany
Quantitative Biology Center (QBiC), University of Tübingen, Germany
article info
Article history:
Received 27 January 2021
Received in revised form
16 March 2021
Accepted 20 March 2021
Available online 26 March 2021
biogenic minerals
iron minerals
Fe(III) reduction
Fe(III) reducing bacteria
Fe(III) minerals play a crucial role for arsenic (As) mobility in aquifers as they usually represent the main
As-bearing phases. Microbial reductive dissolution of As-bearing Fe(III) minerals is responsible for the
release of As and the resulting groundwater contamination in many sites worldwide. So far, in most
studies mainly abiogenic iron minerals have been considered. Yet, biogenic minerals that possess
different properties to their abiogenic counterparts are also present in the environment. In some envi-
ronments they dominate the iron mineral inventory but so far, it is unclear what this means for the As
mobility. We, therefore, performed an in-situ aquifer Fe(III) minerals exposure experiment i) to evaluate
how different biogenic and abiogenic Fe(III) minerals are transformed in a strongly reducing, As-
contaminated aquifer (25 m) compared to As-free moderately reducing aquifer (32 m) and ii) to
assess which microbial taxa are involved in these Fe(III) minerals transformations. We found that higher
numbers of bacteria and archaea were associated with the minerals incubated in the As-contaminated
compared to the non-contaminated aquifer and that all Fe(III) minerals were mainly colonized by
Fe(III)-reducing bacteria, with Geobacter being the most abundant taxon. Additionally, fermenting mi-
croorganisms were abundant on minerals incubated in the As-contaminated aquifer, while methano-
trophs were identied on the minerals incubated in the As-free moderately reducing aquifer, implying
involvement of these microorganisms in Fe(III) reduction. We observed that biogenic Fe(III) minerals
generally tend to become more reduced and when incubated in the As-contaminated aquifer sorbed
more As than the abiogenic ones. Most of abiogenic and biogenic Fe(III) minerals were transformed into
magnetite while biogenic more crystalline mixed phases were not subjected to visible transformation.
This in-situ Fe(III) minerals incubation approach shows that biogenic minerals are more prone to be
colonized by (Fe(III)-reducing) microorganisms and bind more As, although ultimately produce similar
minerals during Fe(III) reduction.
©2021 The Author(s). Published by Elsevier Ltd. This is an open access article under the CC BY license
1. Introduction
Exposure to arsenic (As) is considered a major public health
concern due to its toxic effect on human health and its carcinogenic
potential (National Research Council, 2001). Although As contam-
inates mainly surface water and groundwater, it ultimately enters
the food chain as the water is used for cooking and drinking, while
*Corresponding author. Department of Microbiology, IWWR, Radboud Univer-
sity, the Netherlands.
E-mail address: (M. Glodowska).
AdvectAs Project Membersdlisted in the SI
Contents lists available at ScienceDirect
Environmental Pollution
journal homepage:
0269-7491/©2021 The Author(s). Published by Elsevier Ltd. This is an open access article under the CC BY license (
Environmental Pollution 281 (2021) 117012
anoxic conditions in rice paddy soil lead to the release of As and its
accumulation in crops. (Arain et al., 2009;Molin et al., 2015;
Muhammad et al., 2010). Since As does not change the color, smell
or taste of water and food, it is very difcult to detect its presence
without expensive and time-consuming analysis. In consequence,
many people have been chronically consuming this toxic metalloid
for decades. This is particularly alarming as recent studies have
shown that up to 220 million people worldwide are potentially
exposed to elevated As concentrations in groundwater, overpassing
the WHO limit of 10
g/L (Podgorski and Berg, 2020). What is even
more concerning, the majority of affected areas (94%), where As
concentrations were likely to exceed 10
g/L, were identied in the
most populated regions of Asia. Despite a substantial increase of
awareness regarding the potential risks associated with con-
sumption of As-contaminated water, many people from rural
affected areas simply have no access to safe water or cannot afford
commercially available lters for water purication. Consequently,
As poisoning still remains a global health threat, particularly in the
highly affected areas of South and Southeast Asia.
Although As is an element with minor abundance and an
average concentration of 1.5e3 mg/kg in the continental crust, it is
the 20th most abundant element in the Earths crust (Mandal and
Suzuki, 2002) and it is broadly distributed in the environment
(Lenoble et al., 2004). It was demonstrated that complete mobili-
zation of As in amounts as small as 1 mg/kg from sandy Bangladesh
aquifer sediment would increase the As groundwater concentration
to an extreme value of 7950
g/L (BGS and DPHE, 2001). Previous
research has shown the importance of iron (Fe) minerals, mainly
Fe(III) (oxyhydr)oxides, as main hosting phases for inorganic As,
responsible for As immobilization within sediments (Kontny et al.,
2021;Muehe and Kappler, 2014;Smedley and Kinniburgh, 2002;
Welch et al., 2000). This is very important because on the one hand,
adsorption of As by Fe(III) (oxyhydr)oxides is considered the main
factor controlling As concentrations in the groundwater of many
aquifers (Stollenwerk et al., 2007). On the other hand, one of the
most commonly accepted mechanisms of As mobilization from
sediments to groundwater is via microbially mediated reductive
dissolution of As-bearing Fe(III) minerals (Chatain et al., 2005;Dhar
et al., 2011;Islam et al., 2004;Neumann et al., 2014;Oremland and
Stolz, 2003). Several previous studies have investigated As sorption
capacities of various Fe(III) minerals (e.g., Ona-Nguema et al., 2005).
In particular, high surface area, poorly crystalline Fe(III) phases such
as ferrihydrite (Fh) have a high afnity toward As. Yet, these poorly
crystalline Fe(III) phases are known to be less stable and more
bioavailable as they are preferentially exploited by Fe(III)-reducing
microorganisms as terminal electron acceptor (Lovley et al., 1991).
Therefore, although these poorly crystalline Fe(III) minerals may
sorb more As, they tend to more easily release it under strongly
reducing conditions (Neidhardt et al., 2018).
The high surface area and strong afnity towards metals and
metalloids make Fe(III) minerals an effective sorbent for the
removal of toxic metals from water (Bakatula et al., 2016), and thus
an effective water remediation strategy (Petrusevski et al., 2007). In
consequence, a lot of attention focused on quantifying As sorption
onto different Fe(III) minerals. However, the vast majorityof studies
used abiogenic or synthetic minerals (Dixit and Hering, 2003;
Farquhar et al., 2002;Gimenez et al., 2007;Manning et al., 1998).
Very little work has been done using biogenic Fe(III) minerals
which are produced by microbial activity, either as external or in-
ternal precipitates (Ferris, 2005;Fortin and Langley, 2005;Hao
et al., 2016;Kleinert et al., 2011;Muehe et al., 2016).
This is surprising as microbially driven processes are an integral
part of Fe cycling, contributing to Fe(III) reduction but also to Fe(II)
oxidation and subsequent biogenic Fe(III) mineral formation
(Kappler et al., 2021;Posth et al., 2014). Minerals formed as a result
of direct metabolic activity of bacteria usually occur as nano-
particles or small crystals (2e500 nm) and contain impurities such
as sorbed or co-precipitated Si-ions, PO
, but also cations
such as Mn
, etc. (Fortin and Langley, 2005). Additionally,
these biominerals may contain cell-derived organic matter
(CDOM), and therefore overall possess different properties
compared to their abiogenic counterparts (Hegler et al., 2008;
Muehe et al., 2013b;Posth et al., 2010;Sch
adler et al., 2009).
Nevertheless, biogenic Fe(III) minerals are present in nature and
have been found across different environments (Abramov et al.,
2020;Emerson et al., 1999;Fortin et al., 1998;Ueshima and
Tazaki, 2001). Yet, a limited number of studies has been done us-
ing biogenic Fe(III) minerals which is likely related to the difculty
to synthesize them and the small amounts of minerals that can be
biosynthesized in laboratory setups compared to abiogenic ap-
proaches. A few previous studies have investigated the role of
biogenic Fe(III) minerals in the context of As sorption in the labo-
ratory (Hohmann et al., 2010;Kleinert et al., 2011;Muehe et al.,
2016). However, there is still a lack of knowledge when it comes
to their bioavailability, their efciency in As sorption compared to
abiogenic minerals, their (trans)formation resulting from microbial
reduction, and the microbial community that is involved in their
transformation under environmental conditions. In order to
address these knowledge gaps, we conducted an in-situ incubation
experiment where we incubated a set of biogenic and abiogenic
Fe(III) minerals in a moderately reducing As-free (32 m) and in a
strongly reducing As-contaminated aquifer (25 m) in Vietnam.
2. Materials and methods
2.1. Study area
The study site is located in the village of Van Phuc, about 15 km
SE from Hanoi city, Vietnam, on the banks of the Red River. Several
studies have been conducted at this location, therefore extensive
knowledge about the geology, mineralogy, hydrochemistry and
microbiology is available (Berg et al., 2008;Eiche et al., 2008;
Glodowska et al., 2020c;Kontny et al., 2021;Postma et al., 2017;
Stopelli et al., 2020;van Geen et al., 2013;Wallis et al., 2020). In
brief, the As-contaminated aquifer is of Holocene origin, charac-
terized by strongly reducing conditions, high dissolved As (up to
g/L), Fe (up to 20 mg/L) and CH
(up to 53 mg/L) concentra-
tions. On the contrary, the moderately reducing low-As aquifer is of
Pleistocene origin, which lies adjacent and underneath the Holo-
cene aquifer, has low dissolved As (below 10
g/L), very low Fe
concentrations and not detectable CH
. Between these two aqui-
fers, a sharp transition zone characterized by changing redox con-
ditions was found, where immobilization of As occurs. Due to the
intensive groundwater abstraction in the Hanoi city area and sur-
rounding municipals, a depression cone was formed that led to
inversion of groundwater ow towards the Hanoi city center. As a
consequence, As-contaminated groundwater from the Holocene
aquifer is now owing in the direction of the pristine Pleistocene
aquifer threatening safe water supplies.
2.2. Abiogenic mineral synthesis and sand coating procedure
To coat the quartz (Qtz) grains with Fe minerals, synthetic Fe
minerals (goethite (Gt)-Bayferrox 920, hematite (Hem)-Bayferrox
110, and magnetite (Mt)-Bayferro360 with 0,3
m particle size)
from the company Lanxess/Bayferrox (LANXESS Inorganic Pig-
ments, Krefeld, Germany) were used. All minerals have been tested
with XRD (Bruker D8 Discover; see section 2.6) to ensure that the
manufacturers information is correct. The coating process followed
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
the procedure described previously (Hanna, 2007;Scheidegger
et al., 1993;Schwertmann and Cornell, 2008). To obtain the
coating solution, 100 mg of each mineral was mixed with 10 mL
solution in a 50 mL polyethylene tube and diluted with
(sub-boiled) to adjust the desired pH. In order to obtain the
maximum coating efciency the ionic strength of the NaNO
lution was adjusted to 0.01 M and the pH to 7 for Gt (Schwertmann
and Cornell, 2008), 5 for Mt (Hanna, 2007), and 2.5 for Hem
(Schwertmann and Cornell, 2008). Afterwards the mineral sus-
pensions were shaken at 25
C for 24 h to obtain a homogeneous
suspension. Before mixing with the Fe minerals, the sand was pre-
treated (washed) with 1.0 M HCl solution for 24 h, rinsed several
times with distilled water and dried at 100
C. 2.5 g of the acid-
washed Qtz grains (0.4e0.8 mm, Roth) were added and the
mineral-sand mixture was shaken for another 24 h.
The As content of all reaction components was quantied before
the experiment. Inductively coupled plasma mass spectrometry
(ICP-MS, X-Series 2, ThermoFisher) was used to analyse the initial
solution. The Qtz sand and the synthetic minerals were
analyzed using energy dispersive X-ray uorescence (ED-XRF).
After the mineral coated sand settled, the supernatant was
removed and the coated sand was washed several times with
solution with an ionic strength and pH equal to that of the
reaction medium. Finally, the coated sand was washed free from
any unattached Fe (oxyhydr)oxide rst with a salt solution and then
with pure water using a nylon sieve(63
m). The samples with a pH
above 5 (Gt, Mt) were additionally washed with 1 M NaNO
tion of pH 3 to remove weakly attached Fe (oxyhydr)oxide aggre-
gates. The Gt-, Hem- and Mt-coated sand was nally dried in an
oven at 110
C for 24 h.
Abiogenic 2-line ferrihydrite (Fh) was synthetized following
(Schwertmann and Cornell, 2008), directly in the presence of HCl-
washed Qtz while shaking. The Fh-coated sand was left for 24 h at
room temperature, afterwards washed with MilliQ water in order
to remove any unattached Fe oxyhydroxide particles, and nally
dried at 40
The geochemistry and mineralogy of the Fe-mineral coated sand
grains was determined before and after the incubation experiment.
For more details, see chapter 2.6.
2.3. Biogenic mineral synthesis and sand coating procedure
Biogenic Fh was prepared using the phototrophic Fe(II)-
oxidizing bacterium Rhodobacter ferrooxidans sp. strain SW2 culti-
vated in low-phosphate medium containing 0.14 g KH
, 0.2
NaCl, 0.3 g NH
Cl, 0.5 g MgSO
O, 0.1 g CaCl
O, 1.85 g
per 1 L of MilliQ water and pH adjusted to 7 (Kappler and
Newman, 2004). Medium was prepared anoxically, transferred to
0.5 L Schott bottles and amended with anoxic solutions of 10 mM
and 5 mM acetate. Medium was left for 2 days
in order to allow vivianite and siderite to precipitate. These min-
erals were removed by ltering the medium in the glovebox
through 0.22
m sterile vacuum lter (Millipore, Stericup). In
order to precipitate biogenic Fh directly on the sand, 10 g of HCl-
washed and sterile sand was placed in 150 mL sterile serum bot-
tles, amended with 100 mL of ltered medium, inoculated with 5%
of SW2 culture, closed with a rubber stopper and crimped. All
procedures were performed in the glovebox (100% N
). The head-
space of the serum bottles was exchanged with N
(8:2 ratio). Inoculated serum bottles were kept in horizontal posi-
tion under constant light at 21
C. Bottles were turned every day by
hand within the rst week to ensure homogeneous coating of the
sand. At the end of Fe(II) oxidation, the medium was discarded from
the serum bottle, and the Fh-coated sand was washed with anoxic
MilliQ water several times in order to remove any unattached Fh
and bacterial biomass, and subsequently dried at 40
Biogenic Gt was prepared using the nitrate-reducing Fe(II)-
oxidizing bacterium Acidovorax sp. strain BoFeN1 (Kappler et al.,
2005) and mineral medium containing 0.6 g KH
, 0.3 g NH
0.5 g MgSO
O, 0.1 g CaCl
O and 1.85 g NaHCO
per 1 L
of MilliQ water with pH adjusted to 7.3. Otherwise, Gt-coated sand
was obtained in the same way as described above for biogenic Fh-
coated sand except that inoculated serum bottles were kept in the
dark at 30
Sand coated with a mix of lepidocrocite and goethite (Lep/Gt
) was prepared using Fe(III)-reducing cultures enriched
from sediments collected at our study site in Van Phuc. The cultures
were enriched from sediments collected at 30 m depth with a mix
of acetate and lactate (5 mM each) as electron donor and 2-line
ferrihydrite as electron acceptor under anoxic conditions. The Van
Phuc articial groundwater containing 0.01 g KH
0.025 g MgSO
O, 0.01 g CaCl
O, 0.02 g MnCl
, and
1.85 g NaHCO
per 1 L of MilliQ water with pH adjusted to 7.3 was
used as a growth medium. The culture showed robust Fe(III)-
reducing capacity (see Fig. S1) and Geobacter sp. was determined
as a dominate taxon via denaturing gradient gel electrophoresis
(DGGE) and sequencing of the most prominent bands. However, the
two enrichment cultures had obviously a different microbial com-
munity composition and/or activity since obtained minerals,
although both were identied as a mix of lepidocrocite and
goethite, differed in color, Fe content, and As sorption capacity (see
section 3.1 and 3.3). The abiogenic Fh-coated sand (produced as
described above) and the Van Phuc articial groundwater medium
were inoculated with the enrichment culture and kept in horizontal
position in the dark at 30
C. Bottles were turned every day by hand
until no change of color was observed anymore. The formed
mixture of Lep/Gt was likely a result of secondary mineral forma-
tion during Fh reduction (Boland et al., 2014;Hansel et al., 2005).
Finally, a mix of biogenic lepidocrocite and magnetite (Lep/Mt)
coated sand was prepared as previously described by Sorwat et al.
(2020). The obtained mineral was not pure and consisted of a
mixture of lepidocrocite and magnetite. The identity of the Fe
minerals formed was determined by temperature dependent
magnetic susceptibility.
Again, the geochemistry and mineralogy of the biogenic mineral
coated sand grains was analyzed before and after the incubation
experiment as described below.
2.4. In-situ Fe(III) mineral incubation
The synthesized biogenic and abiogenic Fe(III) minerals coated
on sand were used for the in-situ exposure experiment. One part of
the minerals (biogenic Fh, Gt, Lep/Mt, Lep/Gt
, Lep/Gt
and abio-
genic Fh, Gt, Mt and Hem) was placed in a wafer (a 3D printed frame
covered from both sides with 0.2 mm nylon mesh membrane;
Fig. 1A) as previously described in (Neidhardt et al., 2018). The
second part of the minerals (biogenic Fh, Gt, Lep/Mt, Lep/Gt
abiogenic Fh, Gt, Hem and Mt and uncoated, HCl washed sand used
as a control) was placed directly in pouches made froma permeable
0.2 mm nylon mesh membrane tightened with nylon cable ties
(Fig. S2). Two identical sets of wafers and pouches containing 4.5 g
of each mineral were prepared. Wafers and pouches were attached
to 0.55 mm nylon shing line, weighed down with stainless steel
hex nuts and placed in the monitoring wells for a period of 3 weeks.
One set was incubated in the strongly reducing As-contaminated,
- and CH
-rich well (25 m), while the other one in low As,
and CH
-free well (32 m) (Fig. 1B). Both wells are located
within a few meters from each other (20
and referred to AMS 11(25) and AMS 11(32) in our previous work
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
(Glodowska et al., 2020c;Stopelli et al., 2020) where detailed in-
formation about groundwater hydrochemistry and microbiology
can be found and is summarized in Table S1.
2.5. Sampling and sample preservation
After 3 weeks of incubation, the incubated wafers and pouches
were recovered from the wells. The wafers were immediately
placed in water- and air-tight zip-lock bags ushed with N
kept on dry ice. Bags (PET/PE-LD/Aluminum stand-up pouches
LamiZip, DAKLAPACK) with very low oxygen and water vapor
permeability and protection against UV radiation were used to
minimize sample alteration during transport and storage. The
samples were analyzed for the solid-phase geochemistry and
Minerals incubated in the mesh pouches were immediately
transferred to 15 mL sterile Falcon tubes and immersed in LifeGuard
Soil Preservation Solution (Qiagen) to stabilize microbial DNA and
RNA. Samples were stored on dry ice during transport, placed in
C freezer upon arrival at the laboratory and later used for
16S rRNA gene amplicon sequencing and qPCR analysis.
2.6. Geochemistry and mineralogy
Before analysis, the minerals were removed from the wafer in a
glovebox and dried in a desiccator with drying granulate (Roth
Silica gel orange 1e3 mm) under N
atmosphere (Air liquid UN
1066, 99,8 vol%) at 40
C for 48h to avoid alteration of redox and
temperature sensitive phases. Samples for geochemical analysis
were grinded to a homogenous ne powder (<63
m) using a ball
mill before being analyzed. Fe and As concentrations of grounded
sample material was analyzed using energy dispersive X-ray uo-
rescence (XRF) (Epsilon 5, PANalytical). The ED-XRF detection limit
for As is 1 mg/kg.
Bulk mineral composition of reaction components was deter-
mined by XRD (Bruker D8 Discover) at the LERA (KIT). CuK
ation was used and the samples were scanned from 2
to 82
with an increment of 0,02
at 0.4 s. The relative mineral abun-
dances were estimated from integrated peak areas using the
HighScore-Plus software from Malvern Panalytical GmbH and the
Inorganic Crystal Structure Database (ICSD) for phase identication.
Temperature dependent magnetic susceptibility was determined
on nitrogen ushed samples with a KLY-4S kappabridge (Agico)
from 196
C (liquid nitrogen temperature) to 710
C for Fe mineral
identication. High temperature measurements (room tempera-
ture to 710
C) were measured in an inert argon atmosphere to
minimize oxidation. The analyses were done in the structural ge-
ology division at KIT.
For microscopic analyses, the wafer samples were embedded
directly into a synthetic resin (Araldite, 2020) to avoid any oxida-
tion. The subsequent hardening process was conducted under
vacuum to ensure in-depth impregnation and exclusion of oxygen.
The temperature during the whole process was always below 40
to avoid any thermal alteration of the Fe phases.
The Scanning Electron Microscope (SEM-EDX) measurements
were done using a Tescan Vega 2 SEM with an Oxford Instruments
Silicon Drift Detector (SDD) INCAx-act for Energy Dispersive X-ray
Spectroscopy (EDX) analysis for quantitative elemental analyses
operating at an accelerating voltage of 15 kV, using the secondary
electron (SE) and backscatter (BS) signal. Thin sections were
mounted on specic 3D-printed sample holders and measured
with carbon or gold coatings. Identication of elements in spot
analyses and their distribution using the mapping option of auto-
matic or manual search of elements were performed using the
analytical software SwiftED (version 1.2). Element abundances
Fig. 1. The wafers containing different biogenic (Bio) and abiogenic (Abio) minerals
used for incubation (A). Ferrihydrite (Fh), goethite (Gt), magnetite (Mt), hematite
(Hem), mix of lepidocrocite and goethite (Lep/Gt), and mix of lepidocrocite and
magnetite (Lep/Mt). Schematic cross-section of the As-contaminated strongly reducing
and As-free moderately reducing aquifers with wells where wafers ( ) and mash bags
() containing different minerals were incubated for 3 weeks (B).
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
were determined from the EDS spectra by integrating peak areas
and normalizing the results to 100%. Selected spots were analyzed
with a FEI Quanta 650 FEG ESEM coupled to a Bruker QUANTAX,
Esprit 1.9ED-XRF to obtain higher resolution. All analysis were done
in the Laboratory for Environmental and Raw Material Analysis at
2.7. 16S rRNA gene amplicon sequencing and qPCR
DNA was extracted using a phenol-chloroform method
following a protocol from Lueders et al. (2004). DNA was eluted in
L of a 10 mM Tris buffer. DNA concentration was determined
using a Qubit®2.0 Fluorometer with DNA HS kits (Life Technolo-
gies, Carlsbad, CA, USA). Bacterial and archaeal 16S rRNA genes
were amplied using universal primers 515f: GTGY-
CAGCMGCCGCGGTAA (Parada et al., 2016) and 806r: GGAC-
TACNVGGGTWTCTAAT (Apprill et al., 2015) fused to Illumina
adapters. The PCR cycling conditions were as follows: 95
C for
3 min, 25 cycles of 95
C for 30s, 55
C for 30 s, and 75
C for 30 s.
These steps were followed by a nal elongation step at 72
C for
3 min. The quality and quantity of the puried amplicons were
determined using agarose gel electrophoresis and NanoDrop
(NanoDrop 1000, Thermo Scientic, Waltham, MA, USA). Subse-
quent library preparation steps and Illumina MiSeq sequencing
(Illumina, San Diego, CA, USA) using the 2 250 bp MiSeq Reagent
Kit v2 (500 cycles kit) were performed at Microsynth AG (Balgach,
Switzerland). Between 86,887 and 187,365 read pairs were ob-
tained for each sample.
Sequencing data was analyzed with nf-core/ampliseq v1.1.0,
which includes all analysis steps and software and is publicly
available (Ewels et al., 2020;Straub et al., 2020). Primers were
trimmed, and untrimmed sequences were discarded (<3% per
sample) with Cutadapt version 1.16 (Martin, 2011). Adapter and
primer-free sequences were imported into QIIME2 version 2018.06
(Bolyen et al., 2019), their quality was checked with demux (https://, and they were processed with
DADA2 version 1.6.0 (Callahan et al., 2016) to eliminate PhiX
contamination, trim reads (before median quality drops below 35;
forward reads were trimmed at 230 bp and reverse reads at 219 bp),
correct errors, merge read pairs, and remove PCR chimeras; ulti-
mately, in total 5741 amplicon sequencing variants (ASVs) were
obtained across all samples. Alpha rarefaction curves were pro-
duced with the QIIME2 diversity alpha-rarefaction plugin, which
indicated that the richness of the samples had been fully observed.
A Naive Bayes classier was tted with 16S rRNA gene sequences
extracted from the SILVA version 132 SSU Ref NR 99 database
(Pruesse et al., 2007), using the PCR primer sequences. ASVs were
classied by taxon using the tted classier (
qiime2/q2-feature-classier). ASVs classied as chloroplasts or
mitochondria were removed. The number of removed ASVs was 21,
totaling to <0.5% relative abundance per sample, and the remaining
ASVs had their abundances extracted by feature table (Pruesse
et al., 2007).
Raw sequencing data have been deposited at GenBank under
BioProject accession number PRJNA686278 (https://www.ncbi.nlm.
Quantitative PCRs specic for 16S rRNA genes of bacteria and
archaea, methyl-coenzyme M reductase subunit alpha (mcrA)
genes, particulate methane monooxygenase (pmoA) genes, and
anaerobic arsenite oxidase (arxA) genes were performed. The qPCR
primer sequences, gene-specic plasmid standards, and details on
the thermal programs are given in Table S2. Quantitative PCRs on
DNA extracts obtained as described above were performed in
triplicate using SybrGreen®Supermix (Bio-Rad Laboratories
GmbH, Munich, Germany) on the C1000 Touch thermal cycler
(CFX96real time system). Each quantitative PCR assay was
repeated three times, with triplicate measurements calculated for
each sample per run. Data analysis was done using the Bio-Rad CFX
Maestro 1.1 software version 4.1 (Bio-Rad, 2017).
3. Results and discussion
3.1. Fe(III) mineral dissolution in strongly reducing As-
contaminated and moderately reducing As-free aquifer
The color of most Fe(III) mineral-coated sands (except for
abiogenic Gt, Hem and Mt) used in our experiment became paler
and even faded within 3 weeks of incubation. This implies that
some of the Fe(III) minerals were transformed, dissolved by
reductive dissolution or washed away with the groundwater ow.
The geochemistry provided evidence for the loss of some Fe min-
eral coatings, especially in the Fh and Gt/Lep
samples (Fig. 2A).
We found that in the strongly reducing As-contaminated aquifer
(25 m), in particular the abiogenic and biogenic Fh were reductively
dissolved (loss of Fe
from 1.6 g/kg sand to 0.5 g/kg and from
4.8 g/kg sand to 2.1 g/kg sand for the biogenic and abiogenic Fh,
respectively, i.e., a loss of 67 and 56%) (Fig. 2A). In contrast, the more
crystalline abiogenic Gt, Mt and Hem were not dissolved, indicated
by the rather stable Fe
content of ca. 1.9, 6.0 and 1.2 g/kg sand
before and after incubation (Fig. 2A). Unlike the abiogenic crystal-
line minerals, we found that the biogenic crystalline minerals such
as Gt and the mixture of Lep/Gt
also were reduced and dissolved as
indicated by the change of Fe
concentration from 0.8 to 0.5 g/kg
for Gt and from 14 to 6 g/kg for Lep/Gt
after incubation (loss of 40%
and 58% of Fe
, respectively). Yet, the Lep/Gt
behaved differently
as its Fe
content remained stable (1.7 g/kg) and it was not
When incubating the minerals in the As-free moderately
reducing aquifer (32 m), we observed that Fe
concentration of
poorly crystalline biogenic and abiogenic Fh decreased from 1.6 to
0.3 g/kg and from 4.7 to 3.1 g/kg, respectively, corresponding to a
loss of 83% and 34% of Fe
(Fig. 2A). While more crystalline
abiogenic Fe minerals such as Gt, Mt and Hem remained rather
stable, biogenic Gt, Lep/Gt
, Lep/Gt
and Lep/Mt appeared to be less
stable as indicated by a decrease in the Fe
content of 56, 47, 4
and 11%, respectively (Fig. 2A).
Our observations are supported by previous studies by Zachara
et al. (2002) who found that poorly crystalline Fe(III) phases such as
Fh were unstable under reducing and oxidizing conditions. Further
studies have shown that biominerals are more easily reduced by
microorganisms due to their lower crystallinity, smaller particle
size (larger surface area) and the content of organic molecules
(Langley et al., 2009;Roden, 2004). Hence, biogenic (and abiogenic)
Fe minerals were reduced to certain extent via microbially medi-
ated Fe(III) reductive dissolution within 3 weeks of incubation. It
should be kept in mind however, that biogenic minerals tend to
incorporate and coprecipitate more elements (Al, Ca, Si, P) in the
crystal lattice (Fortin and Langley, 2005;Kiskira et al., 2019;
Voegelin et al., 2010) than abiotic Fe minerals. These incorporations
weaken the crystal structure facilitating its dissolution. Besides,
biogenic Fe minerals such as Mt tend to have smaller grain sizes
than the abiotic counterparts (Karlin, 1990) and a poorly
crystalline-microcrystalline structure which may increase dissolu-
tion of these minerals. Finally, biogenic minerals tend to incorpo-
rate organic substances e.g., bacterial cells (Ferris, 2005) and often
form organo-mineral complexes which can also be dissolved more
easily. Therefore, biogenic minerals are not only preferentially
exploited by microorganisms but are also generally less stable than
abiotically formed Fe minerals which may contribute to more
pronounced loss of Fe in biogenic minerals in our experiment.
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
Although in the uncontaminated aquifer the conditions were less
reducing based on redox potential (Table S1), substantial amounts
of Fe were also mobilized, probably also by microbially mediated
reductive dissolution. It appears that the uncoated control sand
contained some initial coating with a small amount of Fe
(<0.05 wt%) which got dissolved in both highly and moderately
reducing conditions.
Intensive dissolution of Fe(III) minerals was further conrmed
by SEM which revealed that Fe(III) mineral coatings resisted
dissolution mainly in protected areas such as cavities between
closely spaced sand grains (Fig. 3). The rough surfaces of the sand
grains are not uniformly coated by the Fe(III) mineral, either indi-
cating an incomplete original coating of the surface or dissolution.
SEM-EDX spectra of all coatings conrm Fe and O as major ele-
ments and occasionally Si and Al were detected in minor amounts.
Quartz grains coated by abiogenic Fh from the strongly reducing
groundwater (25 m) often show intergrowth of two phases
(Fig. 3A), with a brighter granular phase (Mt) and a gray sur-
rounding phase (Fh). A two-phase intergrowth is also observed for
the biogenic Fh in the moderately reducing groundwater, but
distinctly less obvious (Fig. 3B). The biogenic Lep/Gt is character-
ized by needle-shaped to aky grain shapes and do not show any
transformation (Fig. 3C). Arsenic was not detected using SEM-EDX
due to the high detection limit.
3.2. Fe(III) mineral transformation
Dissolution and transformation are widespread processes
observed in the Fe(III) minerals exposed to As-contaminated
strongly reducing as well as As-free less reducing groundwater
(Figs. 2 and 3). Because of the low concentration of Fe, we were not
able to identify the Fe minerals of the coatings using XRD analysis,
and SEM-EDX provided information about elemental composition
in the minerals only qualitatively. Therefore, we used temperature
dependent magnetic susceptibility in the temperature range
from 190
C to identify the Fe mineralogy and trans-
formation products (Fig. 4). These measurements showed that Mt is
often present in the analyzed samples. Magnetite is a ferrimagnetic
mineral that was identied based on the characteristic Verwey
transition (T
) at about 150
C and its Curie temperature (T
about 580
C. The T
is mostly very weak or even absent, which is
indicative of very small grain size of secondary Mt (original grain
size was 0.3
m) or a signicant vacancy concentration in the
crystal lattice (Dunlop and
Ozdemir, 2007). During the 3 weeks
exposure of Mt in strongly reducing As-rich and less reducing As-
free groundwater, no signicant changes occurred except that the
completely disappeared and T
slightly increased. The relatively
good reversibility of heating and cooling curve and the Hopkinson
peak just before reaching T
indicates pseudo-single domain grain
sizes of magnetite <10
m(Dunlop et al., 2007). The abiogenic
Hem-coated sample shows already little Mt in the original sample.
The higher magnetic susceptibility in the cooling curve indicates
that Hem has transformed into Mt during the high-temperature
measurement in argon atmosphere. In strongly reducing ground-
water (25 m), only Mt remained indicating that Hem is either dis-
solved or transformed into Mt. In moderately reducing
Fig. 2. Changes in Fe (A) and As content (B) in Fe mineral-coated sand after 3 weeks of incubation in As-contaminated strongly reducing water at 25m depth (gray) and As-free
moderately reducing water at 32 m depth (yellow) compared to original minerals before incubation (blue). Ferrihydrite (Fh), goethite (Gt), magnetite (Mt), hematite (Hem), mix of
lepidocrocite and goethite (Lep/Gt), and mix of lepidocrocite and magnetite (Lep/Mt), and uncoated sand used as a control. The % values indicate the loss or gain in Fe content
relative to the initial one before incubation. (For interpretation of the references to color in this gure legend, the reader is referred to the Web version of this article.)
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
Fig. 3. Scanning electron micrographs (backscattering detector) of selected areas of (A) abiogenic Fh-coated sand incubated in the strongly reducing As-contaminated groundwater,
(B) abiogenic Fh-coated sand incubated in the moderately reducing As-free groundwater, and (C) biogenic Lep/Gt
-coated sand incubated in the strongly reducing As-contaminated
groundwater. White squares indicate the analyzed area (zoomed in on the right side). In the right column, the EDX spectra for indicated position (white arrow) is presented.
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
Fig. 4. Temperature depended magnetic susceptibility of (A) abiogenic Mt, (B) abiogenic Hem, (C) abiogenic Fh, (D) biogenic Fh, (E) abiogenic Gt, (F) biogenic Gt, (G) biogenic mix of
and (H) biogenic mix of Lep/Mt before and after incubation in strongly reducing As-contaminated (25 m) and As-free moderately reducing groundwater (32 m). In black:
heating curve, gray: cooling curve.
groundwater (32 m), no change occurred for Hem (Fig. 4B).
Although for the Gt and Fh samples very little Mt in the original
sample cannot be excluded, they are characterized by Mt formation
in both, strongly and moderately reducing groundwater (25 and
32 m) exposure. This seems to be more obvious in the abiogenic
samples of the reduced environment than in the biogenic samples
(Fig. 4CeF). Especially in the Fh samples and in the biogenic Gt the
transformation into Mt is accompanied by strong dissolution
(Fig. 2).
The two biogenic Lep/Gt samples strongly differ in their disso-
lution behavior (Fig. 2) but they both do not show major transitions,
neither in strongly (25 m) nor moderately reducing groundwater
(32 m) (Fig. 4G) suggesting that the biogenic mixed phases are
relatively stable under both of these conditions. This is also the case
for biogenic Lep/Mt although grain size changes in Mt seem to
occur, indicated by the disappearance of the Hopkinson peak in the
heating curves of the groundwater exposed samples (Fig. 4H).
In summary, all samples are characterized by dissolution and/or
transformation processes. The most interesting observation is the
transformation of biogenic and abiogenic Fh and Gt into Mt under
both redox conditions and stability of that mixed biogenic Lep/Gt
and Lep/Mt which are less susceptible for Fe-mineral trans-
formations in these environments (Fig. 5).
3.3. As sorption to Fe minerals
The As concentration in most Fe(III) minerals, incubated in the
strongly reducing water (25 m), is much higher compared to the
Fe(III) minerals that were in contact with As-free moderately
reducing water at 32 m depth (Fig. 2B) which is in agreement with
the previous work at the same site (van Geen et al., 2013). This can
mainly be explained by the higher dissolved As concentration in
groundwater at 25 m depth. Clearly differences in the As content
between the individual Fe(III) minerals (Fig. 2B), however, must be
related to mineral specic characteristics and transformation
processes. The point of zero charge is in a similar range (pH 7.0e7.6)
for all investigated Fe(III) minerals (Borggaard, 1983;Rajput et al.,
2016) and thus, a slightly positive to neutral surface charge is
expectable at the local groundwater pH of ~7.1. This however,
would equally favour the adsorption of As(V), as negatively charged
oxyanion, to all minerals and could not explain the obvious differ-
ences in sorption afnity. The abiogenic Fh-coatings showed by far
the highest sorption capacity towards As (22 mg As/kg), likely
related to its specic surface area which is generally much larger
compared to the other abiogenic Fe(III) minerals that were inves-
tigated (Borggaard, 1983;Rajput et al., 2016). Furthermore (Qi and
Pichler, 2014), have shown that As(III), which is likely the dominant
As species at the redox conditions in the groundwater, clearly
outcompetes As(V) adsorption onto Fh above pH 6. Still, the very
high As concentration of Fh in the strongly reducing groundwater
at 25 m depth is somehow surprising due to the strong dissolution
indicated by decreasing Fe content (Fig. 2A). Incorporation and
xation of As during mineral transformation could be one expla-
nation (Fig. 5).
On the contrary, biogenic Fh did not sorb any detectable As.
Biogenic, poorly crystalline Fe(III) minerals are also known to have a
large specic surface area. However, they have a negative surface
charge at a neutral pH due to the presence of biomass-derived or-
ganics (Posth et al., 2010). This strongly lowers the afnity
regarding binding of As oxyanions. In fact, a previous study from
our lab showed repulsion of negatively charged As(V) by the
negatively charged biogenic Fe(III) (oxyhydr)oxides (Kleinert et al.,
2011). Since water pH under both redox conditions is near neutral
(7.1 and 7.4) low As afnity toward biogenic Fh due to negative
surface charge may be a valid explanation. Additionally to the
change in surface charge by cell-derived organic matter (CDOM)
(Muehe et al., 2013b;Posth et al., 2010), the organic molecules and
other impurities further decrease the sorption capacity by
competitive sorption. Several studies demonstrated that presence
of organic matter generally decrease sorption of As onto Fh (Grafe
Fig. 5. Original Fe(III) minerals and their transformation products after 3 weeks of incubation in the Holocene (25 m) and Pleistocene aquifer (32 m). Ferrihydrite (Fh), goethite (Gt),
magnetite (Mt), hematite (Hem), mix of lepidocrocite and goethite (Lep/Gt), and mix of lepidocrocite and magnetite (Lep/Mt).
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
et al., 2002) and Gt (Grafe et al., 2001).
Surprisingly, the mixed Lep/Gt
minerals biosynthesized by
microbes from Van Phuc sediments sorbed remarkably high and
compared to abiogenic Fh a similar amount of As (20 mg As/kg).
This may be related to a different biomineralization pathway
compared to biogenic Fh. The biogenic Fh was produced as an effect
of microbially mediated precipitation of dissolved Fe
. Previous
studies showed that Fe(III) by-products of this reaction, precipitate
rapidly in the direct vicinity of the cells, leading to its encrustation
(Miot et al., 2009;Sch
adler et al., 2009). Therefore, it is very likely
that biomass incorporated within biogenic Fh prevents As sorption.
On the contrary, the mixed phase of Lep/Gt was produced as an
effect of Fe(III) reduction and secondary mineral precipitation, thus
much less biomass is expected to be incorporated into the freshly
formed mineral. Consequently, this type of mineral is suggested to
sorb As more efciently.
All the other biogenic Fe(III) minerals sorbed ~5 mg As/kg, while
the abiogenic ones (except Fh) sorbed ~2 mg As/kg in the reducing
aquifer. Overall, it can be observed that biogenic Fe(III) minerals
(except Fh) are subjected to much more pronounced dissolution
under strongly and moderately reducing conditions, However, they
tend to sorb more As compared to abiogenic Fe(III) minerals.
3.4. Role of microbial processes in Fe(III) mineral transformation
In order to evaluate the potential involvement of microbial
processes in Fe(III) mineral transformation and reduction, the mi-
crobial community structure and abundance was analyzed in all
samples after 3 weeks of exposure (Figs. 6 and 7). Although,
groundwater hydrochemistry signicantly differed between the
As-contaminated (25 m) and As-free aquifer (32 m), the microbial
community composition from exposed minerals showed certain
3.4.1. Microbial Fe(III) reduction and its inuence on mineral
Putative Fe(III)-reducing bacteria appeared to be the most
abundant group of microorganisms colonizing Fe(III) minerals un-
der both redox-conditions (Fig. 6). Among known Fe(III)-reducers,
taxa related to Geobacter showed the highest relative abundance
in most samples. Geobacter spp. are well-known Fe(III) reducers,
previously shown to be capable of reducing As-bearing Fe(III)
minerals and to contribute to As mobilization (Glodowska et al.,
2020a;Islam et al., 2005a,2005b). These bacteria, besides using
Fe(III) as electron acceptor, may also reduce As(V) while oxidizing
organic carbon (H
ery et al., 2008). Despite its high abundance
among all samples, taxa afliating with Geobacter were mainly
thriving on abiogenic Fh, where they represented 31% and 27% of
total 16S rRNA gene sequences in strongly reducing As-
contaminated (25 m) and less reducing As-free (32 m) ground-
water incubated samples, respectively (Fig. 6). Poorly crystalline Fh
provide the more bioavailable form of Fe(III) that is preferentially
exploited by Fe(III)-reducing microorganisms as a terminal electron
acceptor (Lovley et al., 1991). Moreover, it was previously shown in
batch experiments that Fh is often fully reduced or transformed to
secondary Fe(II)-containing minerals (Fredrickson et al., 1998a,
Fig. 6. Composition and putative function of microbial communities associated with different abiogenic and biogenic Fe(III) minerals, incubated for 3 weeks in an As-contaminated
strongly reducing and in an As-free moderately reducing aquifer. Ferrihydrite (Fh), goethite (Gt), magnetite (Mt), hematite (Hem), mix of lepidocrocite and goethite (Lep/Gt), and
mix of lepidocrocite and magnetite (Lep/Mt).
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
2001), while Gt and Hem are only partially reduced (Zachara et al.,
1998). This might explain more advanced dissolution of biogenic
and abiogenic Fh in both environments, compared to Hem and Gt.
Furthermore, in our experiment, the more crystalline minerals such
as Hem and Gt revealed substantially lower relative abundance of
Geobacter and other Fe(III) reducers (3e16%), suggesting a generally
lower bioavailability of these minerals. The Holophagaceae family
was the second most abundant taxon found on incubated Fe(III)
minerals, however, many sequences within this family remained
unassigned. Further analysis of these unassigned sequences using
BLAST revealed that most of the Holophagaceae taxa were closely
related to Geothrix fermentans, another well-known Fe(III) reducer
(Coates et al., 1999;Islam et al., 2005a), which was probably
involved in magnetite formation, as previously demonstrated
(Klueglein et al., 2013). Taxa afliating with Rhodocyclaceae,Bur-
kholderiaceae,Myxococcales,Clostridiaceae and Deferrisoma that
were previously shown or suggested to be involved in Fe(III)
reduction (Li and Peng, 2011;P
erez-Rodríguez et al., 2016;
Slobodkina et al., 2012;Treude et al., 2003;Zhang et al., 2020), were
also abundant among all analyzed mineral samples. Overall, it ap-
pears that the relative abundance of Fe(III)-reducing taxa was
higher in biogenic minerals incubated in the moderately reducing
(32 m) rather than in those incubated in strongly reducing
groundwater (25 m). Yet, bacterial and archaeal 16S rRNA gene
copy number was higher in Fe(III) minerals retrieved from As-
contaminated aquifer (25 m), implying that generally more mi-
croorganisms thrive in As-, Fe- and CH
-rich water (Fig. 7). Besides
abiogenic Fh that hosted the highest relative abundance of Fe(III)
reducers, it can be noticed that biogenic Fe(III) minerals tended to
host more bacteria and archaea (Fig. 7) and showed higher relative
abundance of putative Fe(III)-reducers compared to abiogenic
minerals (Fig. 6). This indicates that biogenic Fe(III) minerals will
likely get reduced or transformed by dissimilatory metal-reducing
bacteria faster than abiogenic minerals and might explain why
biogenic minerals were generally more dissolved compared to
abiogenic ones.
The end product of the microbially mediated Fe(III) reduction
depends on several factors. Besides the identity of primary min-
erals, the identity of secondary minerals depends on the medium
composition, concentration of electron donor and acceptor, Fe(III)
mineral age, sorbed ions, co-associated Fe(III) oxides (Zachara et al.,
2002) and likely on the abundance and the composition of the
microbial community mediating the reaction. Therefore, it is very
difcult to predict what kind of secondary minerals might be
produced under natural conditions. It was however shown that Fh,
if not completely reduced, is often transformed into secondary
Fe(II)-containing minerals, such as siderite (in buffered medium),
magnetite or green rust (Fredrickson et al., 1998b;Muehe et al.,
2013a,2013b). More crystalline minerals such as Gt and Hem
tend to remain largely unchanged or get partially converted to
vivianite, magnetite or green rust (Cutting et al., 2009;Muehe et al.,
2013b;Zachara et al., 1998). This is in agreement with our results
(Fig. 5) which showed almost complete dissolution of Fh or its
partial transformation to Mt (mainly under strongly reducing
condition) and only fractional transformation of Gt and Hem.
3.4.2. The putative role of fermenters in Fe(III) reduction
Based on 16S rRNA gene sequencing (Fig. 6), it appears that
typical fermenters belonging to Firmicutes (Acidaminobacter,Syn-
trophomonadaceae), Bacteroidetes (Lentimicrobiaceae,Prolixibacter-
aceae), Chloroexi (Anaerolineaceae) and Acidobacteria
(Aminicenantales)(Gupta et al., 2014;Kampmann et al., 2012;Tan
et al., 2015;Wang et al., 2020) were abundant among exposed
minerals. These fermenters were particularly abundant on minerals
incubated in the strongly reducing shallow aquifer (25 m) where
they accounted for 8e14% of the microbial community. This is likely
related to the fact, that groundwater at this depth had 4 times more
dissolved organic carbon (DOC) compared to the deeper moder-
ately reducing aquifer (Table S1). Previous studies already indicated
that many fermenters can use Fe(III) as electron acceptor (Dong
et al., 2017;Esther et al., 2015;Zhao et al., 2020). Therefore, we
hypothesize that fermenters abundantly found in Fe(III) minerals
incubated in the As-contaminated strongly reducing aquifer (25 m)
were in fact exploiting Fe(III) minerals as electron acceptors and
subsequently contributed to Fe(III) reduction while oxidizing DOC.
The relative abundance of putative fermenters appeared to be
similar in biogenic and abiogenic minerals incubated in the As-
contaminated aquifer (25 m). On the contrary, only biogenic Fh
and to a lower extent biogenic Gt, incubated in the As-free envi-
ronment (32 m), revealed a higher relative abundance of fermen-
ters, which represented 14% and 4% of the microbial community,
respectively. Our previous study investigating microbial commu-
nities across both aquifers in Van Phuc revealed that fermenters
were indeed a dominant microbial group in groundwater
(Glodowska et al., 2020c). Despite their dominance among exposed
minerals, Fe(III) reducers, although present in most of the wells,
represented only a small fraction (up to 2%) of the groundwater
microbial community. This implies that Fe(III) reducers require
physical contact with minerals in order to use Fe(III) as electron
acceptor, while fermenters may use a wide range of electron
Fig. 7. Bacterial and archaeal 16S rRNA gene, pmoA and mcrA copy numbers per 1 g
mineral coated sand in biomass associated with abiogenic and biogenic Fe(III) minerals
that were incubated for 3 weeks in an As-contaminated strongly reducing and an As-
free moderately reducing aquifer. Ferrihydrite (Fh), goethite (Gt), magnetite (Mt), he-
matite (Hem), mix of lepidocrocite and goethite (Lep/Gt), and mix of lepidocrocite and
magnetite (Lep/Mt).
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
acceptors present in the water.
3.4.3. Involvement of methanotrophs in Fe(III) reduction and
Methanotrophs are another microbial group enriched on the
exposed minerals (Fig. 6). Interestingly, these microorganisms were
highly abundant on minerals incubated in the lower As-free aquifer
(32 m) where CH
is undetectable, while they were much less
abundant in the strongly reducing upper aquifer (25 m) which is
characterized by a very high CH
concentration (51 mg/L; Table S1).
We have observed this phenomenon also in our previous study in
Van Phuc, where methanotrophs appeared to be much more
abundant in the CH
-free groundwater compared to groundwater
rich in CH
(Glodowska et al., 2020c) and in the deeper sediments
associated with CH
-free groundwater compared to shallow sedi-
ments where CH
-rich groundwater is present (Glodowska et al.,
2021). We hypothesize that this might be due to 1) intensive
methanotrophy that led to depletion of CH
in the deeper part of
the aquifer and to signicant biomass production or 2) shortage of
electron acceptors in the strongly reduced aquifer. Either way,
methanotrophs were apparently attracted to Fe(III) minerals
exposed to uncontaminated moderately reducing water (32 m) as
they represented between 4 and 21% of the microbial community.
On the contrary, methanotrophs represented only 1e4% of 16S
rRNA gene sequences in the minerals incubated in the strongly
reducing groundwater (25 m). Several previous studies have
demonstrated that anaerobic CH
oxidation can be coupled to
Fe(III) reduction (Aromokeye et al., 2020;Cai et al., 2018;Ettwig
et al., 2016). Moreover, our earlier study conrmed that this pro-
cess indeed takes place in Van Phuc and contributes to As mobili-
zation (Glodowska et al., 2020b). However, in all of the previous
studies archaea belonging to Methanosarcinales were shown to
catalyze this process and to date there was no report of bacteria
capable of mediating this reaction. Yet, the methanotrophic com-
munity thriving on the Fe(III) minerals in our experiment was
generally dominated by bacteria related to Candidatus Methyl-
omirabilis,Methyloparacoccus and Crenothrix. Abiogenic crystalline
Fe(III) minerals such as Hem, Mt and Gt, incubated in less reducing
groundwater (32 m) appeared to be preferentially associated with
Ca. Methylomirabilis which represented 21, 19 and 17% of the mi-
crobial community, respectively. Although we previously found this
taxon mainly in the moderately reducing As-free environment, its
abundance was never that high, neither in the groundwater (max.
8%) nor in the sediments (max. 5%) (Glodowska et al., 2020b;
Glodowska et al., 2021). A previous study already suggested that Ca.
Methylomirabilis can use Fe(III) instead of NO
as electron acceptor
(Shen et al., 2019). Therefore it can be hypothesized that bacteria
belonging to Ca. Methylomirabilis can use Fe(III) as electron acceptor
either alone or in syntropic collaboration with other community
members. The Fe(III) minerals exposed to As-contaminated
strongly reducing groundwater (25 m) were much less associated
with methanotrophs and taxa related to Methyloparacoccus and
Crenothrix being most abundant. Here, also abiogenic more crys-
talline minerals such as Mt and Gt were preferentially colonized.
The high abundance of pmoA and mcrA genes among the microbial
community associated with Fe(III) minerals further highlight the
involvement of methanotrophs in Fe(III) reduction and trans-
formation (Fig. 7).
4. Conclusions
Biogenic Fe(III) minerals have different properties compared to
their abiogenic counterpart and, in consequence, they behave
differently in the natural environment. Our study indicates that
biogenic minerals are preferentially occupied by bacteria and host
higher relative abundances of Fe(III) reducers. As a result, this type
of Fe(III) minerals are more prone to transformation and reduction
even under moderately reducing conditions. Geobacter and Hol-
ophagaceae taxa are probably key players in Fe(III) reduction in
both aquifers. Furthermore, the results of our exposure experiment
implies involvement of fermenters and methanotrophs in Fe(III)
The environmental conditions are important factors inuencing
the behavior of Fe(III) minerals and their stability. At the investi-
gated eld site we observed a strong reductive dissolution of
biogenic Fe(III) minerals or their transformation into Mt. On the
other hand, biogenic minerals (except for Fh) tend to sorb more As.
If As is sorbed/incorporated into Mt, into further secondary trans-
formation products or into the original Fe(III) minerals has to be
investigated in future by
-XANES studies. Biominerals such as Lep/
Gt biosynthesized using native microbial communities from Van
Phuc were less prone to reduction and transformation and showed
much higher afnity towards As, suggesting that this type of min-
erals might be more representative for this specic environment. It
also implies that As and Fe dynamic in Van Phuc is largely
controlled by indigenous microbial community.
Moreover, webelieve that using a set of different Fe(III) minerals
as a decoy may be an efcient strategy to assess the Fe(III) reducing
capacity of the native microbial community and to capture the key
Fe(III) reducers for further study and isolation.
Credit author statement
Martyna Glodowska: Conceptualization, Methodology, Investi-
gation, Writing eoriginal draft, Writing ereview &editing, Visu-
alization. Magnus Schneider: Conceptualization, Methodology,
Writing ereview &editing, Investigation, Visualization. Elisabeth
Eiche: Writing ereview &editing, Formal analysis. Agnes Kontny:
Writing ereview &editing, Formal analysis. Thomas Neumann:
Supervision, Funding acquisition, Project administration. Daniel
Straub: Data curation, Formal analysis, Software. Sara Kleindienst:
Writing ereview &editing, Funding acquisition, Resources.
Andreas Kappler: Supervision, Writing ereview &editing, Funding
acquisition, Resources
Declaration of competing interest
The authors declare that they have no known competing
nancial interests or personal relationships that could have
appeared to inuence the work reported in this paper.
The authors thank all AdvectAs project members for the
collaboration and support. Special thanks to Pham Hung Viet, Pham
Thi Kim Trang, Vi Mai Lan, Mai Tran and Viet Nga from Hanoi
University of Science for the assistance during the sampling
campaign. This study was supported by the Deutsche For-
schungsgemeinschaft (DFG) (KA 1736/41e1). D. Straub is funded by
the Institutional Strategy of the University of Tübingen (DFG, ZUK
63) and by the Collaborative Research Center CAMPOS (Grant
Agreement SFB 1253/1 2017). S. Kleindienst is funded by an Emmy-
Noether fellowship (DFG, grant #326028733). The authors
acknowledge support by the High Performance and Cloud
Computing Group at the Zentrum für Datenverarbeitung of the
University of Tübingen, the state of Baden-Württemberg through
bwHPC and the German Research Foundation (DFG) through grant
no INST 37/935-1 FUGG.
M. Glodowska, M. Schneider, E. Eiche et al. Environmental Pollution 281 (2021) 117012
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... The total relative abundance of this azoreductase-producing group in Phases 1, 2, and 3 were 3.1%, 6.7%, and 8.3%, respectively. They are also known as dissimilatory iron-reducing bacteria [17,20]. In addition, Anaerolineaceae, Burkholderiales, and Syntrophobacter were discovered and enriched with the participation of Fe 3 O 4 . ...
In this study, ferroferric oxide (Fe3O4) was used to accelerate treatment performance of synthetic wastewater containing azo dye reactive black and starch as a sole carbon source. In a batch experiment, 10 g·L⁻¹ of Fe3O4 was confirmed to be optimal to this type of wastewater treatment. Therefore, this concentration was chosen for the follow-up experiment with a continuous system of 10 L anaerobic baffled reactor (ABR). The ABR was run at 30 ℃ with a hydraulic retention time of 18.6 h for 170 days. Fe3O4 was added to each ABR compartment at 10 g·L⁻¹ from day 105. The Fe3O4 supplement significantly enhanced ABR treatment performance. With influent color and chemical oxygen demand (COD) levels of 1300 Pt-Co and 305 mgCOD·L⁻¹, their removal efficiencies increased remarkably from 44% and 42% in the Fe3O4-free phase to 85% and 76% in the Fe3O4-added phase, respectively. Interestingly, the ratio of generated methane on removed COD in Fe3O4-involved phases were lower than that in Fe3O4-free phase, suggesting a preference of the ABR for decolorization over methanogenesis in an iron-rich condition. Microbial community analysis revealed that Fe3O4 may have immensely contributed to dyeing degradation process through the selective enrichment of key substrate-fermenting and dye-degrading bacteria, especially Clostridium spp.
... Several studies have shown a coupling between Fe-oxidation and Fe-reduction in the environment, i.e. situations in which biogenic Fe(III) minerals produced by Fe-oxidizing bacteria are used as substrates for Fe-reducers [Blöthe andRoden, 2009, Emerson et al., 2010]. Interestingly, experiments have shown that biogenic Fe(III) minerals are more easily colonized and reduced than abiogenic ones [e.g., Glodowska et al., 2021], suggesting that Fe-reducers may rely on Fe-oxidizers to supply bioavailable Fe(III) substrates in some environments. Such cooperative metabolic relationships may be particularly relevant to environments where Fe-reduction and Fe-oxidation processes spatially overlap [e.g., Laufer et al., 2016]. ...
... Several studies have shown a coupling between Fe-oxidation and Fe-reduction in the environment, i.e. situations in which biogenic Fe(III) minerals produced by Fe-oxidizing bacteria are used as substrates for Fe-reducers [Blöthe andRoden, 2009, Emerson et al., 2010]. Interestingly, experiments have shown that biogenic Fe(III) minerals are more easily colonized and reduced than abiogenic ones [e.g., Glodowska et al., 2021], suggesting that Fe-reducers may rely on Fe-oxidizers to supply bioavailable Fe(III) substrates in some environments. Such cooperative metabolic relationships may be particularly relevant to environments where Fe-reduction and Fe-oxidation processes spatially overlap [e.g., Laufer et al., 2016]. ...
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Prokaryotes have been shaping the surface of the Earth and impacting geochemical cycles for the past four billion years. Biomineralization, the capacity to form minerals, is a key process by which microbes interact with their environment. While we keep improving our understanding of the mechanisms of this process (“how?”), questions around its functions and adaptive roles (“why?”) have been less intensively investigated. Here, we discuss biomineral functions for several examples of prokaryotic biomineralization systems, and propose a roadmap for the study of microbial biomineralization through the lens of adaptation. We also discuss emerging questions around the potential roles of biomineralization in microbial cooperation and as important components of biofilm architectures. We call for a shift of focus from mechanistic to adaptive aspects of biomineralization, in order to gain a deeper comprehension of how microbial communities function in nature, and improve our understanding of life co-evolution with its mineral environment.
... Huyen et al. [48] observed that As release may be controlled by sorptiondesorption reactions with clays/phyllosilicates, where As is in competition for sorption sites with In situ incubated experiments in aquifers have revealed the elevated efficiency of biogenic and poorly crystalline Fe(III) minerals (i.e. ferrihydrite) in As sorption compared to abiogenic minerals and their preeminent availability to more easily release As in dissolved form under reducing conditions by (Fe(III)-reducing) microorganisms [72]. ...
Among potential toxic elements (PTEs), arsenic (As) and mercury (Hg) are well known for the toxicity of their different chemical species and diffusion in the environment via several anthropic sources (i.e., industrial settlements, mining activity). Bottom sediments often become a repository for As and Hg, although they may be considered a potential secondary source of these elements into the water column depending on their speciation and mobility. Focusing on the most recent studies (since 2019) on the occurrence of As and Hg in contaminated aquatic sediments, the aim of this review is to give an overview of the current understandings on the complex biogeochemical cycle of these elements in this environmental media. The main biogeochemical factors governing the transformations of As and Hg among their different chemical species were synthesised, highlighting those driving the formation of more mobile and/or bioavailable forms. Additionally, the most advanced analytical techniques for the determination of the different chemical species of As and Hg in sediments are briefly presented.
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The drinking water quality of millions of people in South and Southeast Asia is at risk due to arsenic (As) contamination of groundwater and insufficient access to water treatment facilities. Intensive use of nitrogen (N) fertilizer increases the possibility of nitrate (NO 3 ⁻ ) leaching into aquifers, yet very little is known about how the N cycle will interact with and affect the iron (Fe) and As mobility in aquifers. We hypothesized that input of NO 3 ⁻ into highly methanogenic aquifers can stimulate nitrate-dependent anaerobic methane oxidation (N-DAMO) and subsequently help to remove NO 3 ⁻ and decrease CH 4 emission. We, therefore, investigated the effects of N input into aquifers and its effect on Fe and As mobility, by running a set of microcosm experiments using aquifer sediment from Van Phuc, Vietnam supplemented with ¹⁵ NO 3 ⁻ and ¹³ CH 4 . Additionally, we assessed the effect of N-DAMO by inoculating the sediment with two different N-DAMO enrichment cultures (N-DAMO(O) and N-DAMO(V)). We found that native microbial communities and both N-DAMO enrichments could efficiently consume nearly 5 mM NO 3 ⁻ in 5 days. In an uninoculated setup, NO 3 ⁻ was preferentially used over Fe(III) as electron acceptor and consequently inhibited Fe(III) reduction and As mobilization. The addition of N-DAMO(O) and N-DAMO(V) enrichment cultures led to substantial Fe(III) reduction followed by the release of Fe ²⁺ (0.190±0.002 mM and 0.350±0.007 mM, respectively) and buildup of sedimentary Fe(II) (11.20±0.20 mM and 10.91±0.47 mM, respectively) at the end of the experiment (day 64). Only in the N-DAMO(O) inoculated setup, As was mobilized (27.1±10.8 μg/L), while in the setup inoculated with N-DAMO(V) a significant amount of Mn (24.15±0.41 mg/L) was released to the water. Methane oxidation and ¹³ CO 2 formation were observed only in the inoculated setups, suggesting that the native microbial community did not have sufficient potential for N-DAMO. An increase of NH 4 ⁺ implied that dissimilatory nitrate reduction to ammonium (DNRA) took place in both inoculated setups. The archaeal community in all treatments was dominated by Ca . Methanoperedens while the bacterial community consisted largely of various denitrifiers. Overall, our results suggest that input of N fertilizers to the aquifer decreases As mobility and that CH 4 cannot serve as an electron donor for the native NO 3 ⁻ reducing community. Graphical abstract
The coprecipitation of magnetite (Fe3O4) with arsenic (As) is a potential remediation technique for As-contaminated groundwater that can be applied to meet increasingly stringent As drinking water limits. However, knowledge of the fate of As coprecipitated with magnetite during aging for extended periods is lacking, which is critical to predict the long-term efficiency of this As treatment strategy. In this work, I combined aqueous As measurements with solid-phase characterization by synchrotron-based Fe and As K-edge X-ray absorption spectroscopy (XAS) to track the transformation of magnetite and the speciation of coprecipitated As(V) or As(III) for up to a year in oxic or anoxic conditions. It was determined that the initial magnetite particle increased in crystallinity for all aging experiments, but some differences in solid-phase Fe speciation were detected depending on aging conditions. For the anoxic aging samples with initial As(V), a significant fraction (15% of the total Fe) of maghemite (a magnetic Fe oxide spinel with formulγ-Fe2O3) was identified, which was coupled to As(V) reduction [As(III) was â∼30% of the total sorbed As], suggesting electron transfer between magnetite and particle-bound As(V). In the oxic aging experiments, the initial particle crystallized, with a large fraction of Fe(III) (oxyhydr)oxides (i.e., maghemite and lepidocrocite, γ-FeOOH) in the final products. Despite increased crystallinity suggested by Fe XAS analysis, sorbed As was not released from the particles in any experiment (aqueous As never exceeded 1 μg/L). This remarkable stability of As coprecipitated with magnetite was revealed by As K-edge XAS to be largely due to the formation of distinct multinuclear As uptake modes [i.e., As(V) incorporation; hexanuclear 3C As(III) complexes]. These results demonstrate the unique potential of magnetite for long-term As sequestration.
Copper oxide nanoparticles (CuO NPs) are widely used as fungicides in agriculture. Arsenic (As) is a ubiquitous contaminant in paddy soil. The present study was focused on the adsorption behavior of CuO NPs with regard to As as well as the characteristics of the microbial community changes in As-contaminated soil-rice systems in response to CuO NPs. The study found that CuO NPs could be a temporary sink of As in soil; a high dose of CuO NPs promoted the release of As from crystalline iron oxide, which increased the As content in the liquid phase. The study also found that the As bioavailability changed significantly when the dose of CuO NPs was higher than 50 mg kg-1 in the soil-rice system. The addition of 100 mg kg-1 CuO NPs increased the microbial diversity and the abundance of genes involved in As cycling, decreased the abundance of Fe(III)-reducing bacteria and sulfate-reducing genes, and decreased As accumulation in grains. Treatment with 500 mg kg-1 CuO NPs increased the abundance of Fe(III)-reducing bacteria and sulfate-reducing genes, decreased Fe plaques, and increased As accumulation in rice. The adverse effects of CuO NPs on crops and associated risks need to be considered carefully.
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Arsenic groundwater contamination threatens the health of millions of people worldwide, particularly in South and Southeast Asia. In most cases, the release of arsenic from sediment was caused by microbial reductive dissolution of arsenic-bearing iron(III) minerals with organic carbon being used as microbial electron donor. Although in many arsenic-contaminated aquifers high concentrations of methane were observed, its role in arsenic mobilization is unknown. Here, using microcosms experiments and hydrogeochemical and microbial community analyses, we demonstrate that methane functions as electron donor for methanotrophs, triggering the reductive dissolution of arsenic-bearing iron(III) minerals, increasing the abundance of genes related to methane oxidation, and ultimately mobilizing arsenic into the water. Our findings provide evidence for a methane-mediated mechanism for arsenic mobilization that is distinct from previously described pathways. Taking this together with the common presence of methane in arsenic-contaminated aquifers, we suggest that this methane-driven arsenic mobilization may contribute to arsenic contamination of groundwater on a global scale.
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One of the major methods to identify microbial community composition, to unravel microbial population dynamics, and to explore microbial diversity in environmental samples is high-throughput DNA- or RNA-based 16S rRNA (gene) amplicon sequencing in combination with bioinformatics analyses. However, focusing on environmental samples from contrasting habitats, it was not systematically evaluated (i) which analysis methods provide results that reflect reality most accurately, (ii) how the interpretations of microbial community studies are biased by different analysis methods and (iii) if the most optimal analysis workflow can be implemented in an easy-to-use pipeline. Here, we compared the performance of 16S rRNA (gene) amplicon sequencing analysis tools (i.e., Mothur, QIIME1, QIIME2, and MEGAN) using three mock datasets with known microbial community composition that differed in sequencing quality, species number and abundance distribution (i.e., even or uneven), and phylogenetic diversity (i.e., closely related or well-separated amplicon sequences). Our results showed that QIIME2 outcompeted all other investigated tools in sequence recovery (>10 times fewer false positives), taxonomic assignments (>22% better F-score) and diversity estimates (>5% better assessment), suggesting that this approach is able to reflect the in situ microbial community most accurately. Further analysis of 24 environmental datasets obtained from four contrasting terrestrial and freshwater sites revealed dramatic differences in the resulting microbial community composition for all pipelines at genus level. For instance, at the investigated river water sites Sphaerotilus was only reported when using QIIME1 (8% abundance) and Agitococcus with QIIME1 or QIIME2 (2 or 3% abundance, respectively), but both genera remained undetected when analyzed with Mothur or MEGAN. Since these abundant taxa probably have implications for important biogeochemical cycles (e.g., nitrate and sulfate reduction) at these sites, their detection and semi-quantitative enumeration is crucial for valid interpretations. A high-performance computing conformant workflow was constructed to allow FAIR (Findable, Accessible, Interoperable, and Re-usable) 16S rRNA (gene) amplicon sequence analysis starting from raw sequence files, using the most optimal methods identified in our study. Our presented workflow should be considered for future studies, thereby facilitating the analysis of high-throughput 16S rRNA (gene) sequencing data substantially, while maximizing reliability and confidence in microbial community data analysis.
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The fate of arsenic (As) in groundwater is determined by multiple interrelated microbial and abiotic processes that contribute to As (im)mobilization. Most studies to date have investigated individual processes related to As (im)mobilization rather than the complex networks present in situ. In this study, we used RNA-based microbial community analysis in combination with groundwater hydrogeochemical measurements to elucidate the behavior of As along a 2 km transect near Hanoi, Vietnam. The transect stretches from the riverbank across a strongly reducing and As-contaminated Holocene aquifer, followed by a redox transition zone (RTZ) and a Pleistocene aquifer, at which As concentrations are low. Our analyses revealed fermentation and methanogenesis as important processes providing electron donors, fueling the microbially mediated reductive dissolution of As-bearing Fe(III) minerals and ultimately promoting As mobilization. As a consequence of high CH4 concentrations, methanotrophs thrive across the Holocene aquifer and the redox transition zone. Finally, our results underline the role of SO42--reducing and putative Fe(II)-/As(III)-oxidizing bacteria as a sink for As, particularly at the RTZ. Overall, our results suggest that a complex network of microbial and biogeochemical processes has to be considered to better understand the biogeochemical behavior of As in groundwater.
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Geogenic groundwater arsenic (As) contamination is pervasive in many aquifers in south and southeast Asia. It is feared that recent increases in groundwater abstractions could induce the migration of high-As groundwaters into previously As-safe aquifers. Here we study an As-contaminated aquifer in Van Phuc, Vietnam, located ~10 km southeast of Hanoi on the banks of the Red River, which is affected by large-scale groundwater abstraction. We used numerical model simulations to integrate the groundwater flow and biogeochemical reaction processes at the aquifer scale, constrained by detailed hydraulic, environmental tracer, hydrochemical and mineralogical data. Our simulations provide a mechanistic reconstruction of the anthropogenically induced spatiotemporal variations in groundwater flow and biogeochemical dynamics and determine the evolution of the migration rate and mass balance of As over several decades. We found that the riverbed–aquifer interface constitutes a biogeochemical reaction hotspot that acts as the main source of elevated As concentrations. We show that a sustained As release relies on regular replenishment of river muds rich in labile organic matter and reactive iron oxides and that pumping-induced groundwater flow may facilitate As migration over distances of several kilometres into adjacent aquifers.
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Natural organic matter (NOM) can contribute to arsenic (As) mobilization as an electron donor for microbially-mediated reductive dissolution of As-bearing Fe(III) (oxyhydr)oxides. However, to investigate this process, instead of using NOM, most laboratory studies used simple fatty acids or sugars, often at relatively high concentrations. To investigate the role of relevant C sources, we therefore extracted in situ NOM from the upper aquitard (clayey silt) and lower sandy aquifer sediments in Van Phuc (Hanoi area, Vietnam), characterized its composition, and used 100-day microcosm experiments to determine the effect of in situ OM on Fe(III) mineral reduction, As mobilization, and microbial community composition. We found that OM extracted from the clayey silt (OMC) aquitard resembles young, not fully degraded plant-related material, while OM from the sandy sediments (OMS) is more bioavailable and related to microbial biomass. Although all microcosms were amended with the same amount of C (12 mg C/L), the extent of Fe(III) reduction after 100 days was the highest with acetate/lactate (43 ± 3.5% of total Fe present in the sediments) followed by OMS (28 ± 0.3%) and OMC (19 ± 0.8%). Initial Fe(III) reduction rates were also higher with acetate/lactate (0.53 mg Fe(II) in 6 days) than with OMS and OMC (0.18 and 0.08 mg Fe(II) in 6 days, respectively). Although initially more dissolved As was detected in the acetate/lactate setups, after 100 days, higher concentrations of As (8.3 ± 0.3 and 8.8 ± 0.8 μg As/L) were reached in OMC and OMS, respectively, compared to acetate/lactate-amended setups (6.3 ± 0.7 μg As/L). 16S rRNA amplicon sequence analyses revealed that acetate/lactate mainly enriched Geobacter, while in situ OM supported growth and activity of a more diverse microbial community. Our results suggest that although the in situ NOM is less efficient in stimulating microbial Fe(III) reduction than highly bioavailable acetate/lactate, it ultimately has the potential to mobilize the same amount or even more As.
High arsenic (As) concentrations in groundwater are a worldwide problem threatening the health of millions of people. Microbial processes are central in the (trans)formation of the As-bearing ferric and ferrous minerals, and thus regulate dissolved As levels in many aquifers. Mineralogy, microbiology and dissolved As levels can vary sharply within aquifers, making high-resolution measurements particularly valuable in understanding the linkages between them. We conducted a high spatial resolution geomicrobiological study in combination with analysis of sediment chemistry and mineralogy in an alluvial aquifer system affected by geogenic As in the Red River delta in Vietnam. Microbial community analysis revealed a dominance of fermenters, methanogens and methanotrophs whereas sediment mineralogy along a 46 m deep core showed a diversity of Fe minerals including poorly crystalline Fe (II/III) and Fe(III) (oxyhydr)oxides such as goethite, hematite, and magnetite, but also the presence of Fe(II)-bearing carbonates and sulfides which likely formed as a result of microbially driven organic carbon (OC) degradation. A potential important role of methane (CH4) as electron donor for reductive Fe mineral (trans)formation was supported by the high abundance of Candidatus Methanoperedens, a known Fe(III)-reducing methanotroph. Overall, these results imply that OC turnover including fermentation, methanogenesis and CH4 oxidation are important mechanisms leading to Fe mineral (trans)formation, dissolution and precipitation, and thus indirectly affecting As mobility by changing the Fe-mineral inventory.
Biogeochemical cycling of iron is crucial to many environmental processes, such as ocean productivity, carbon storage, greenhouse gas emissions and the fate of nutrients, toxic metals and metalloids. Knowledge of the underlying processes involved in iron cycling has accelerated in recent years along with appreciation of the complex network of biotic and abiotic reactions dictating the speciation, mobility and reactivity of iron in the environment. Recent studies have provided insights into novel processes in the biogeochemical iron cycle such as microbial ammonium oxidation and methane oxidation coupled to Fe(iii) reduction. They have also revealed that processes in the biogeochemical iron cycle spatially overlap and may compete with each other, and that oxidation and reduction of iron occur cyclically or simultaneously in many environments. This Review discusses these advances with particular focus on their environmental consequences, including the formation of greenhouse gases and the fate of nutrients and contaminants.
Iron minerals are the most important arsenic host in As-contaminated deltaic sediments. Arsenic release from Fe minerals to groundwater exposes millions of people worldwide to a severe health threat. To understand the coupling of Fe mineralogy with As (im)mobilization dynamics, we analyzed the geochemistry and mineralogy of a 46 m long sediment core drilled into the redox transition zone where a high As Holocene aquifer is juxtaposed to a low As Pleistocene aquifer in the Red River delta, Vietnam. We specifically concentrated on mm- to cm-scale redox interfaces within the sandy aquifer. Various Fe phases, such as Fe- and Mn- bearing carbonates, pyrite, magnetite, hematite and Fe-hydroxides (goethite, lepidocrocite) with distinct As concentrations were identified by a combination of high-resolution microscopic, magnetic and spectroscopic methods. The concentration of As and its redox species in the different Fe-minerals were quantified by microprobe analysis and synchrotron X-ray absorption. We developed a conceptual model integrating Fe-mineral transformations and related As (im)mobilization across the redox interfaces. Accordingly, As is first mobilized via the methanogenic dissolution of Fe(III) (oxyhydr)oxide mineral coatings on sand grains when reducing groundwater from the Holocene aquifer intruded into the Pleistocene sands. This stage is followed by the formation of secondary Fe(II)-containing precipitates (mainly Fe- and Mn-bearing carbonates with relatively low As < 70 μg/g), and minor pyrite (with high As up to 5800 μg/g). Due to small-scale changing redox conditions these Fe(II) minerals dissolve again and the oxidative behavior of residual Fe(III)-phases in contact with the reducing water leads to the formation of abundant Fe(III)/Fe(II) (oxyhydr)oxides especially at the studied redox interfaces. Microcrystalline coatings and cementations of goethite, magnetite and hematite have intermediate to high As sorption capacity (As up to 270 μg/g) creating a key sorbent responsible for As (im)mobilization at interfingering redox fronts. Our observations suggest a dynamic system at the redox interfaces with coupled redox reactions of abiotic and biotic origin on a mm to cm-scale. In a final stage, further reduction creates magnetite with low As sorption capacity as important secondary Fe-mineral remaining in reduced gray Pleistocene aquifer sands while considerable Fe and As is released into the groundwater. The presented redox-dependent sequence of Fe phases at redox interfaces provides new insights of their role in As (im)mobilization in reducing aquifers of south and southeast Asia.
Magnetite nanoparticles are often promoted as remediation agents for heavy metals such as chromium due to their reactivity and high surface area. However, their small size also makes them highly mobile increasing the risk that reacted pollutants will be transported to different locations rather than being safely controlled. Released to aquatic environments, aggregation leads to a loss of their nano-specific properties and contaminant-removal capacity. We immobilized magnetite onto sand to overcome these issues whilst maintaining reactivity. We compare biogenic magnetite and abiogenic magnetite coated sand against magnetite nanoparticles. Magnetite coatings mostly exhibited a Fe(III)/Fe(III) ratio close to stoichiometry (0.5). We tested the efficacy of the magnetite-coated sand to adsorb chromium, with respect to biogenic/abiogenic nanoparticles. Langmuir-type sorption of Cr(VI) onto magnetite (4.32 mM total Fe) was observed over the tested concentration range (10–1000 µM). Biogenic nanoparticles showed the highest potential for Cr(VI) removal with maximum adsorption capacity (Qmax) of 1250 µmol Cr/g Fe followed by abiogenic nanoparticles with 693 µmol Cr/g Fe. All magnetite coated sands exhibited similar sorption behavior with average Qmax ranging between 257−471 µmol Cr/g Fe. These results indicate coating magnetite onto sand may be more suitable than free nanoparticles for treating environmental pollutants such as chromium.
Dowsing for danger Arsenic is a metabolic poison that is present in minute quantities in most rock materials and, under certain natural conditions, can accumulate in aquifers and cause adverse health effects. Podgorski and Berg used measurements of arsenic in groundwater from ∼80 previous studies to train a machine-learning model with globally continuous predictor variables, including climate, soil, and topography (see the Perspective by Zheng). The output global map reveals the potential for hazard from arsenic contamination in groundwater, even in many places where there are sparse or no reported measurements. The highest-risk regions include areas of southern and central Asia and South America. Understanding arsenic hazard is especially essential in areas facing current or future water insecurity. Science , this issue p. 845 ; see also p. 818