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Revision as of 20:05, 14 April 2021

Passive Sampling of Sediments

"Passive sampling" refers to a group of methods used to quantify the availability of organic contaminants to move between different media and/or to react in environmental systems such as indoor air, lake waters, or contaminated sediment beds. To do this, the passive sampling material is deployed in the environmental system and allowed to absorb chemicals of interest via diffusive transfers from the surroundings. Upon recovery of the passive sampler, the accumulated contaminants are measured, and the concentrations in the sampler are interpreted to infer the chemical concentrations in specific surrounding media like porewater in a sediment bed. Such data are then useful inputs for site assessments such as those seeking to quantify fluxes from contaminated sediment beds to overlying waters or to evaluate the risk of significant uptake into benthic infauna and the larger food web.

Related Article(s):

Contributor(s): Dr. Philip M. Gschwend

Key Resource(s):

  • Validating the Use of Performance Reference Compounds in Passive Samplers to Assess Porewater Concentrations in Sediment Beds[1]
  • In situ passive sampling of sediments in the Lower Duwamish Waterway Superfund site: Replicability, comparison with ex situ measurements, and use of data[2]
  • Laboratory, Field, and Analytical Procedures for Using Passive Sampling in the Evaluation of Contaminated Sediments: User’s Manual[3]

Introduction

Environmental media such as sediments typically contain many different materials or phases, including liquid solutions (e.g. water, nonaqueous phase liquidslike spilled oils) and diverse solids (e.g., quartz, aluminosilicate clays, and combustion-derived soots). Further, the chemical concentration in the porewater medium includes both molecules that are "truly dissolved" in the water and others that are associated with colloids in the porewater[4][5][6]. As a result, contaminant chemicals distribute among these diverse media (Figure 1) according to their affinity for each and the amount of each phase in the system[7][8][9][10][11]. As such, the chemical concentration in any one medium (e.g., truly dissolved in porewater) in a multi-material system like sediment is very hard to know from measures of the total sediment concentration, which unfortunately is the information typically found by analyzing for chemicals in sediment samples.

If an animal moves into this system, it will also accumulate the chemical in its tissues from the loads in all the other materials (Figure 1). This is important if one is concerned with exposures of the chemical to other organisms, including humans, who may eat such shellfish. Predicting the quantity of contaminant in the clam requires knowledge of the relative affinities of the chemical for the clam versus the sediment materials. For example, if one knew the chemical's truly dissolved concentration in the porewater and could reasonably assume the chemical of interest in the clams has mostly accumulated in its lipids (as is often the case for very hydrophobic compounds), then one could estimate the chemical concentration in the clam (Cclam, typically in units of μg/kg clam wet weight) using a lipid-water partition coefficient, Klipid-water, typically in units of (μg/kg lipid)/(μg/L water), and the porewater concentration of the chemical (Cporewater, in μg/L) with Equation 1.

Equation 1. Cclam = flipid x Klipid-water x Cporewater
where:
flipid is the fraction lipids contribute to the total wet weight of a clam (kg lipid/kg clam wet weight), and
Cporewater is the freely dissolved contaminant concentration in the porewater surrounding the clam.

While there is a great deal of information on the values of Klipid-water for many chemicals[12], it is often very inaccurate to estimate truly dissolved porewater concentrations from total sediment concentrations using assumptions about the affinity of those chemicals for the solids in the system[7]. Further, it is difficult to isolate porewater without colloids and/or measure the very low truly dissolved concentrations of hydrophobic contaminants of concern like polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), nonionic pesticides like dichlorodiphenyltrichloroethane (DDT), and polychlorinated dibenzo-p-dioxins (PCDDs)/ dibenzofurans (PCDFs)[13].

Passive Samplers

One approach to address this problem for contaminated sediments is to insert organic polymers like low density polyethylene (LDPE), polydimethylsiloxane (PDMS), or polyoxymethylene (POM) that can absorb such chemicals in the sediment[14][15][16][17][18][19][2]. In this approach, the polymer is inserted in the sediment bed where it absorbs some of the contaminant load via the contaminant's diffusion into the polymer from the surroundings. When the polymer achieves sorptive equilibration with the sediments, the chemical concentration in the polymer, Cpolymer (μg/kg polymer), can be used to find the corresponding concentration in the porewater, Cporewater (μg/L), using a polymer-water partition coefficient, Kpolymer-water ((μg/kg polymer)/(μg/L water)), that has previously been found in laboratory testing[20][21], as shown in Equation 2.

         Equation 2. Cporewater = Cpolymer / Kpolymer-water

Such “passive uptake” by the polymer also reflects the availability of the chemicals for transport to adjacent systems (e.g., overlying surface waters) and for uptake into organisms (e.g., bioaccumulation). Thus, one can use the porewater concentrations to estimate the biotic accumulation of the chemicals, too. For example, for the concentration in the clam equilibrated with the sediment, Cclam (μg/kg clam), would be found by combining Equations 1 and 2 to get Equation 3.

         Equation 3. Cclam = flipid x Klipid-water x Cpolymer / Kpolymer-water

Performance Reference Compounds (PRCs)

Perhaps unsurprisingly, pollutants with low water solubility like PAHs, PCBs, etc. do not diffuse quickly through sediment beds. As a result, their accumulation in polymeric materials in sediments can take a long time to achieve equilibration[22][23][24]. This problem was recognized previously for passive samplers called semipermeable membrane devices (SPMDs, e.g. polyethylene bags filled with triolein[25]) that were deployed in surface waters. As a result, representative chemicals called performance reference compound (PRCs) were dosed inside the samplers before their deployment in the environment, and the PRCs' diffusive losses out of the SPMD could be used to quantify the fractional approach toward sampler-environmental surroundings equilibration[26][25]. A similar approach can be used for polymers inserted in sediment beds[22][1]. Commonly, isotopically labeled forms of the compounds of interest such as deuterated or 13C-labelled PAHs or PCBs are homogeneously impregnated into the polymers before their deployments. Upon insertion of the polymer into the sediment bed (or overlying waters or even air), the initially evenly distributed PRCs begin to diffuse out of the sampling polymer and into the sediment (Figure 2).

Assuming the contaminants of interest undergo the same mass transfer restrictions limiting their rates of uptake into the polymer (e.g., diffusion through the sedimentary porous medium) that are also limiting transfers of the PRCs out of the polymer[22][1], then fractional losses of the PRCs during a particular deployment can be used to adjust the accumulated contaminant loads to what they would have been at equilibrium with their surroundings with Equation 4.

Equation 4. C(∞)polymer = C(t)polymer / fPRC lost
where:
fPRC lost is the fraction of the PRC lost to outward diffusion,
C(∞)polymer is the concentration of the contaminant in the polymer at equilibrium, and
C(t)polymer is the concentration of the contaminant in the polymer after deployment time, t.                

Since investigators are commonly interested in many chemicals at the same time, it is impractical to have a PRC for each contaminant of interest. Instead, a representative set of PRCs is used to characterize the rates of polymer-environment exchange as a function of the PRCs' properties (e.g., diffusivities, partition coefficients), the sediments characteristics (e.g., porosity), and the nature of the polymer used (e.g., film thickness, affinity for the chemicals)[22][23]. The resulting mass transfer model fit can then be used to estimate the fractional approaches to equilibrium for many other contaminants, whose diffusive and partitioning properties are also known. And these fractions can be used to adjust the target chemical concentrations that have accumulated from the sediment into the same polymeric sampler to find the equilibrated results[1]. Finally, these equilibrated concentrations can be used in Eq. 2 to estimate truly dissolved contaminant concentrations in the sediment's porewater.

Field Applications

Polymeric materials can be deployed in sediment in various ways[3]. PDMS coatings on silica rods, called SPMEs (solid phase micro extraction devices), can be incorporated into slotted rods, while thin sheets of polymers like LDPE or POM can be incorporated into sheet metal frames. In both cases, such hardware is used to insert the polymers into sediment beds (Figure 3).

Deployment of the assembled passive samplers can be done via poles from a boat[1], by divers[2], or by attaching the samplers to a sampling platform lowered off a vessel[27]. Typically, the method used depends on the water depth. Small buoys on short lines, sometimes with associated water-sampling polymeric materials in mesh bags (see right panel of Figure 3), are attached to the samplers to facilitate the sampler recoveries. After recovery, the samplers are wiped to remove any adhering sediment, biofilm, or precipitates and returned to the laboratory for PRC and target contaminant analyses. The resulting measurements of the accumulated target chemical concentrations can be adjusted using the observed PRC losses and publicly available software programs



Figure 1. Conceptual model of mercury speciation in the environment[28]

Mercury (Hg) is released into the environment typically in the inorganic form. Natural emissions of Hg(0) come mainly from volcanoes and the ocean. Anthropogenic emissions are mainly from artisanal and small-scale gold mining, coal combustion, and various industrial processes that use Hg ( see the UN Global mercury assessment). Industrial and natural emissions of gaseous elemental mercury, Hg(0), can travel long distances in the atmosphere before being oxidized and deposited on land and in water as inorganic Hg(II). The long range transport and atmospheric deposition of Hg results in widespread low-level Hg contamination of soils at concentrations of 0.01 to 0.3 mg/kg[29].

Hg-contaminated sites are most commonly contaminated with Hg(II) from industrial discharge and have soil concentrations in the range of 100s to 1000s of mg/kg[29]. Direct exposure to Hg(II) and Hg(0) can be a human health risk at heavily contaminated sites. However, the organic form of Hg, methylmercury (MeHg or CH3Hg+) is typically the greater concern. MeHg is a neurotoxin that is particularly harmful to developing fetuses and young children. Direct contamination of the environment with MeHg is not common, but has occurred, most notably in Minamata Bay, Japan (see also Minamata disease). More commonly, MeHg is formed in the environment from Hg(II) in oxygen-limited conditions in a processes mediated by anaerobic microorganisms. Because MeHg biomagnifies in the aquatic food web, MeHg concentrations in fish can be elevated in areas that have relatively low levels of Hg contamination. The MeHg production depends heavily on site geochemistry, and high total Hg sediment concentrations do not always correlate with MeHg production potential.

Biogeochemistry/Mobility of Hg in soils

In the environment, Hg mobility is largely controlled by chelation with various ligands or adsorption to particles[30]. Hg(II) is most strongly attracted to the sulfur functional groups in dissolved organic matter (DOM) and to sulfur ligands. Over time, newly released Hg(II) “ages” and becomes less reactive to ligands and is less likely to be found in the dissolved phase. Legacy Hg(II) found in sediments and soils is more likely to be strongly adsorbed to the soil matrix and not very bioavailable compared to newly released Hg(II)[30]. MeHg has mobility tendencies similar to Hg, with DOM and sulfur ligands competing with each other to form complexes with MeHg[31]. However, unlike Hg-S complexes, MeHg-S does not have limited solubility.

The bioavailability of Hg(II) is one of the factors controlling MeHg production in the environment. MeHg production occurs in anoxic environments and is affected by: (1) the bioavailability of Hg(II) complexes to Hg- methylating microorganisms, (2) the activity of Hg-methylating microorganisms, and (3) the rate of biotic and abiotic demethylation. MeHg is produced by anaerobic microorganisms that contain the hgcAB gene[32]. These microorganisms are a diverse group and include, sulfate-reducing bacteria, iron-reducing bacteria, and methanogenic bacteria. Site geochemistry has a significant effect on MeHg production. Methylating microorganisms are sensitive to oxygen, and MeHg production occurs in oxygen-depleted or anaerobic zones in the environment, such as anoxic aquatic sediments, saturated soils, and biofilms with anoxic microenvironments[33]. The activity of methylating microorganisms can be impacted by redox conditions, the concentrations of organic carbon, and different electron acceptors (e.g. sulfate vs iron)[33]. Overall, MeHg concentrations and production are impacted by demethylation as well. Demethylation can occur both abiotically and biotically and occurs at a much faster rate than methylation. The main routes of abiotic demethylation are photochemical reactions and demethylation catalyzed by reduced sulfur surfaces[34][35]. Methylmercury can be degraded biotically by aerobic bacteria containing the mercury detoxification, mer operon and through oxidative demethylation by anaerobic microorganisms[34].

Bioaccumulation and Toxicology

Regulatory criteria are most often based on total Hg concentrations, however, MeHg is the form of Hg that can bioaccumulate in wildlife and is the greatest human and ecological health risk[36]. MeHg represents over 95% of the Hg found in fish[37]. Hg and MeHg can be taken up directly from contaminated water into organisms, with the identity of the Hg-ligand complexes determining how readily the Hg is taken up into the organism[38]. Direct bioconcentration from water is the major uptake route at the base of the food web. Hg and MeHg can also enter the food web when benthic organisms ingest contaminated sediments[39]. Further up the food web organisms are exposed to Hg and MeHg both through exposure to contaminated water and through their diet. The higher up the trophic level, the more important dietary exposure becomes. Fish obtain more than 90% of Hg from their diet[38].

Humans are mainly exposed to Hg in the forms of MeHg and Hg(0). Hg(0) exposure comes from dental amalgams and industrial/contaminated site exposures. Hg(0) readily crosses the blood/brain barrier and mainly effects the nervous system and the kidneys[40]. MeHg exposure comes from the consumption of contaminated fish. In the human body, MeHg is readily absorbed through the gastrointestinal tract into the bloodstream and crosses the blood/brain barrier, affecting the central nervous system. MeHg can also pass through the placenta to the fetus and is particularly harmful to the developing nervous system of the fetus.

MeHg and Hg toxicity in the body occurs through multiple pathways and may be linked to the affinity of Hg for sulfur groups. Hg and MeHg bind to S-containing groups, which can block normal bodily functions[41].

Regulatory Framework for Mercury

In the United States, mercury is regulated by several different environmental laws including: the Mercury Export Ban Act of 2008, the Mercury-Containing and Rechargeable Battery Management Act of 1996, the Clean Air Act, the Clean Water Act, the Emergency Planning and Community Right-to-Know Act, the Resource Conservation and Recovery Act, and the Safe Drinking Water Act[42].

In 2013, the United States signed the international Minamata Convention on Mercury. The Minamata Convention on Mercury seeks to address and reduce human activities that are contributing to widespread mercury pollution. Worldwide, 128 countries have signed the Convention.

Remediation Technologies

As a chemical element, Hg cannot be destroyed, so the goal of Hg-remediation is immobilization and prevention of food web bioaccumulation. At very highly contaminated sites (>100s ppm), sediments are often removed and landfilled[29]. In situ capping is also a common remediation approach. Both dredging and capping can be costly and ecologically destructive, and the development of less invasive, less costly remediation technologies for Hg and MeHg contaminated sediments is an active research field. Eckley et al.[29]and Wang et al.[43] give thorough reviews of standard and emerging technologies.

Recently application of in situ sorbents has garnered interest as a remediation solution for Hg[29]. Many different materials, including biochar and various formulations of activated carbon, are successful in lowering porewater concentrations of Hg and MeHg in contaminated sediments[44]. More research is needed to determine whether Hg and MeHg sorbed to these materials are available for uptake into organisms. Site biogeochemistry can also impact the efficacy of sorbent materials, with dissolved organic matter and sulfide concentrations impacting Hg and MeHg sorption. Overall, knowing site biogeochemical characteristics is important for predicting Hg mobility and MeHg production risks as well as for designing a remediation strategy that will be effective.

References

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  2. ^ 2.0 2.1 2.2 Apell, J.N., and Gschwend, P.M., 2016. In situ passive sampling of sediments in the Lower Duwamish Waterway Superfund site: Replicability, comparison with ex situ measurements, and use of data. Environmental Pollution, 218, pp. 95-101. DOI: 10.1016/j.envpol.2016.08.023   Authors’ Manuscript
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See Also

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