H e a l t h a s p e c t s
Pathogens and micro pollutants
Occurrence, fate and behaviour
of pharmaceutical active compounds (PAC)
and endocrine disrupting compounds (EDC)
Persistent organic substances
Abstract
A simple nomogram was developed to describe the number of log-removals of pathogens and biodegradable organics between injection and recovery wells for ASR with nearby production wells, for simple Aquifer Storage Transfer and Recovery (ASTR) projects, and between the river bank and production wells in bank filtration projects. The method assumes a homogeneous isotropic aquifer of uniform thickness and porosity with uniform ambient hydraulic gradient, uniform rate of pumping and extraction, a constant exponential rate of biodegra- dation or pathogen inactivation, and linear adsorption isotherm. Only two non-dimensional parameters were needed to define the number of log10reductions or biodegradation during transport through the aquifer to the recovery well. These describe advective transport due to pumping wells and due to the regional hydraulic gradient respectively. These parameters uniquely defined the ratio of minimum travel time to the time for one-log10 reduction, and therefore defined the number of log10 reductions of the contaminant reaching neighbouring wells.
The method was applied to a case study in the Bandung Basin, Indonesia to derive safe distances between stormwater injection wells and downgradient pumping wells for nine hydrogeological zones in the basin.
Recharge enhancement was being considered to help address the groundwater overdraft in the basin. While the approach is very simple, it was adequate to demonstrate that in each of these zones the rate of groundwater flow was high in relation to rate of pathogen inactivation, so that runoff from the ground surface, where stormwater and sewage systems merged, should not be admitted into wells. However roof-runoff, piped directly to a well, was likely to yield water suitable for injecting into the aquifer without adverse impacts on the quality of neighbouring wells. Although the model is very simple, it may be useful as a planning tool or to assist in designing MAR projects.
Keywords
Groundwater, solute transport, analytical models, pathogens, ASR, Indonesia.
I N T R O D U C T I O N
A simple screening tool was developed to evaluate the potential for viable pathogens to be transported from an injection well to a down-gradient pumping well. Through the use of common simplifying assumptions, such as steady-state flow and a constant exponential rate of pathogen attenuation, the model could be reduced to a nomo- gram to represent a wide array of situations. The method was applied to the Bandung Basin to assess the water qual- ity impacts of injecting stormwater and roof-runoff into the aquifer.
for managed recharge of aquifers
Peter Dillon, Paul Pavelic, Karen Barry,
Susanne Fildebrandt and Notoadmodjo Prawoto
M E T H O D
Firstly it is assumed that the pathogens (or other trace organic contaminants) undergo first-order exponential decay with respect to residence time in the aquifer.
Eqn. (1) where Ctis the concentration or number of viable pathogens per unit volume after storage time t, Cois the con- centration or number in water recharged via the injection well, and τis the time required for the initial concentra- tion or number to be reduced to 10% of its original value, often called the one-log10removal time.
S
Siin ng glle e w we ellll ssyysstte em mss ((A AS SR R))
The minimum residence time of injected water in the aquifer in ASR systems is simply the storage period between injection and recovery, defined here as ts. Hence the worst case scenario for biodegradation or inactivation is when t = ts in equation (1), neglecting the potential effect of dilution with native groundwater.
D
Du ua all w we ellll ssyysstte em mss
In Aquifer Storage Transfer and Recovery (ASTR) or with ASR and a nearby pumping well which needs water quality to be protected, the worst-case scenario considers the water that has travelled to the recovery well along the shortest flow path when the injection and recovery wells are operating continuously (at the same rate). This gives the minimum travel time (tmin) over which biodegradation can occur. If the ambient groundwater velocity is small with respect to the gradients induced by the injection and recovery wells, and assuming injection and pumping rates are equal;
Eqn. (2)
where: D is the aquifer thickness (m);
L is the distance between injection and recovery wells (m);
ne is the porosity of the aquifer (–);
Q is the rate of steady-state pumping (in and out) (m3d–1).
The shortest travel time occurs when there is a regional hydraulic gradient in the aquifer and the recovery well is situated directly downgradient of the injection well. Dilution with ambient groundwater is neglected as this effect leads to enhanced contaminant attenuation. In such a two well system, tminis given by Rhebergen and Dillon, (1999) as:
Eqn. (3)
where νdois the Darcian velocity component between the injection well to the recovery well (md–1). As continuous concurrent injection and recovery rarely occur, this equation is likely to underestimate travel time because when wells are operated intermittently the average hydraulic effective gradient over the time of travel will be less than the value which has been assumed in this equation (worst-case scenario).
Assuming that the contaminant or pathogen of interest is also sorbed onto the aquifer matrix with a linear adsorp-
do e
L v D
Q L t n
+
= 3
min
π Q
L n
t D e
3
2 min
= π
τ / 0 10 t
t C
C = −
T O P I C 4
Health aspects / Pathogens and micro pollutants 478
tion isotherm, then contaminant transport is slowed by a constant retardation factor, R, with respect to conservative transport of the water molecules. Thus the minimum travel time, t min iof a contaminant, i, is a factor R times the travel time of conservative solutes that move at the same rate as the water, ie:
t min i= R tmin Eqn. (4)
where R = 1 + Kdρ/ ne Eqn. (5)
and Kd= focKoc Eqn. (6)
where Kdis the distribution coefficient for a linear adsorption isotherm [m3/kg];
ρis the dry bulk density of the porous media [kg m–3];
focis the weight fraction of organic carbon in the porous media [–];
Kocis the adsorption coefficient related to organic carbon content [m3/ kg OC].
In this model, sorption acts only to extend the travel time during which biodegradation takes place, and by itself is not regarded as a sustainable attenuation process. It is also assumed that all of the water from the pumping well originates from the injection well.
B
Ba an nk k ffiillttrra attiio on n
From image well theory, constant head along a stream can be approximated by an injection and recovery well pair bisected by the stream, as in equation (2), but substituting the distance, a, between the stream and the pumping well (a = L /2) and noting that the water derived from the stream has only half the travel distance of the water travelling from the image well to the pumping well. This results in equation (7) (Dillon et al., 2002).
Eqn. (7)
In this instance the rate of streambed infiltration, q, induced by pumping from the well at a rate, Q, at any time, t, since the commencement of pumping, is approximated by Glover and Balmer (1954) as:
Eqn. (8)
Where α is aquifer diffusivity (transmissivity /storage coefficient) and here in an unconfined system of approx- imately constant saturated thickness, D, and hydraulic conductivity, K, is KD /ne. For steady state pumping, the value of q /Q is one for a semi-infinite aquifer with an initially horizontal free-surface.
A contaminant originating from the stream at a steady concentration, C0, reaches a concentration, Ct in the well at time t is given by Equation (9) which is a restatement of equation (1) accounting for equations, (4), (7) and (8).
Dillon et al. (2002) also account for the effect of conservative contaminants, e.g. salinity, in groundwater for bank filtration design.
Eqn. (9)
τ / 0 10 Rtmin
t Q
C q
C = −
⎟⎠
⎜ ⎞
⎝
= ⎛
t erfc a Q
q
4α Q
a n
t D e
3
2 2
min
= π
O
Otth he err ssiittu ua attiio on nss
Wherever the travel time between a constant sole source of contaminant and a point of groundwater discharge can be calculated, along with the relative contribution of that source to the discharge, and retardation and degradation or inactivation rates are known, then the concentration remaining at the point of discharge may be calculated.
ASRRI (Aquifer Storage and Recovery Risk Index) is a computer program recently developed by CSIRO that, using the principles outlined above, calculates the risk of contamination for a range of single and dual well ASR systems (Miller et al., 2002). The risk indexes calculate whether specific trace organic or microbial contaminants in the recovered water would reach its target attenuation ratio (log removal) or its guideline value for a given scenario.
N O M O G R A M F O R D UA L W E L L S Y S T E M
Using three nondimensional terms, each with a physical meaning, a nomogram based on equations 1, 3 and 4 was produced to define the log10 removal under a range of scenarios for dual well systems. This allows separation distances or pumping rates to be determined that will meet any required number of log-removals for a given set of aquifer characteristics (depth, porosity, Darcian velocity and retardation factor (organic carbon fraction). The non- dimensional terms are as follows:
Term 1: Advective transport expression due to pumping wells ;
Term 2: Advective transport due to regional hydraulic gradient ;
Term 3: The ratio of contaminant travel time to one log10removal time .
The nomogram, shown in Figure 1, can be used in various ways. Knowing existing or planned well locations and aquifer properties, and an approximate one-log10removal time, e.g. from Dillon et al. (2005), a value for term 1 can be found on the x axis and a curve value for term 2 can be used to interpolate a y coordinate value on the nomogram. This immediately defines the value of term 3 which also defines the log10reduction expected for that
τ tmin doτ
e
v R L n
τ
2
Q R D L ne T O P I C 4
Health aspects / Pathogens and micro pollutants 480
0 2 4 6
0 5 10 15 20
(neL2DR)/(Q )
log removal
100 20 10 5 2 neLR / v
Figure 1. Nomogram for dual well ASR system (recovery well down-gradient of injection well)
contaminant. Alternatively, if the aim is to achieve a given log removal (say 4-log) then the curve parameter (term 2) cannot be less than 5. The curve parameter (term 2) is composed of variables that are intrinsic to the aquifer and the contaminant or pathogen, and the sole control variable is the separation distance, L. Hence L can, in principle, be increased so that it is sufficiently large that the curve rises above the number of log removals required.
Term 1 contains two control variables, and is proportional to L squared and inversely proportional to Q. In this case if Term 2 had a value of 10, then Term 1 would need a value exceeding 7 to achieve the required 4 log removals.
This may require reducing the pumping rate.
The nomogram shows that there are some combinations of characteristics where no realistic system design will produce sufficient log removal for satisfactory performance. In those areas ASR should be prevented unless the injected water is adequately pre-treated. Conversely, higher ne, L and R, values increase terms 1 and 2 to give more time available for biodegradation, and result in a higher log removal. Note that the curve parameter tends towards infinity as the ambient gradient approaches zero. Fildebrandt et al. (2003) also provide nomograms for recovery wells positioned upgradient of the injection well and at various angles, however normally the down gradient direction is expected to be the most critical case for decision making concerning water quality impacts on the aquifer.
A P P L I C AT I O N T O B A N D U N G B A S I N, I N D O N E S I A
The city of Bandung in western Java, Indonesia, is facing increasing stress on the quantity and quality of its ground- water resources due to the rising population, increasing water demand, and changes of land use. Groundwater abstraction in the Bandung Basin increased from 47 Mm3 to 61 Mm3per year between 1990 and 1994, while the number of deep wells increased from 1,000 to 4,700 (Suryantoro, 1999) and groundwater levels continued to fall by 2 – 4 m per year (Soetrisno 1998). In recent years the Indonesian government has become more active in ground- water protection and has considered developing methods to enhance groundwater recharge (Bukit 1995; Supriyo et al., 1999), including use of excess stormwater runoff.
The Bandung Basin is surrounded by a number of volcanic mountains in the north and south. The Citarum River originates in the south and flows northwest through a central 1,000 km2 plain where most of the urban and industrial areas are located. The climate is tropical with an annual rainfall of 1,900 –2,200 mm with distinct wet and dry seasons (Soetrisno, 1998). The central basin is composed of various Quaternary fluvial sediments of volcanic origin. The general direction of groundwater flow is from the northern and southern mountain ranges towards the Citarum River. There are two hydrogeological systems: the shallow aquifers and the deep aquifers. The shallow Kosambi aquifer occurs within the upper 40 m. and has low (< 2 L /s) or low to moderate well yields (2 –10 L /s) and was originally of potable quality (Suhari and Siebenhuner, 1993). The deep aquifersare semi-confined to con- fined and are present at depths of between 40 m and 150 m. The two uppermost deep aquifers are the Cibeurum aquifer and the underlying Cikapundung aquifer. The Cibeurum aquifer, composed of young volcanic deposits from which most water is pumped, has moderate to high well yields (2 –10 L /s to >10 L /s) whilst the Cikapundung aquifer, which consists of old volcanic deposits, has moderate well yields (2 –10 L /s) (Soetrisno, 1998).
The fate of pathogens in the groundwater presents the greatest single threat to public health from recharge enhan- cement using waters of impaired quality. Therefore this study aimed to find separation distances between injection wells and pumping wells down-gradient that would achieve specified levels of microbial attenuation in different zones of the Bandung Basin. For illustration purposes here, a 4-log removal is suggested for injected surface water whilst for inherently better quality roof-runoff a 1-log removal may be enough.
The shallow aquifers of the Bandung Basin were classified into regions with different hydraulic gradients (<1%, 1–10% and 10 –17%) and well yield (<2, 2 –10, and >10 L /s) (Suhari and Siebenhuner, 1993). These maps were
overlain to form zones to allow the assignment of L, the distance between the injection and recovery well necessary for a certain log removal. There were nine possible combinations (3 hydraulic gradients x 3 well yields). Zone 1 produced the smallest separation distance and zone 9 the largest. Two of the nine different zones (7 and 9) did not occur in the basin (Figure 2). Zones 1, 2 and 4 occupy the largest the area.
To calculate L values for each zone the following parameter values were assigned: K = 1 m /d, ne = 0.1, D = 10 m, R = 1 (i.e. no retardation) and assuming τ= 10 days (Toze 2004). Sensitivity analyses are reported in Fildebrandt et al. (2003). The calculated separation distances (L) for 1-log removal are 45 to 150 m, and for 4-log removal are 92 to 326 m (Table 1).
Table 1. Recommended separation distances (m) between injection and recovery wells for different zones for 1-log removal and 4-log removal, ττ= 10 days
τ= 10 d Estd. yield and Darcian flow D = 10 m, ne = 0.1
Zone Q [l /s] Vdo [m /d] 1-log 4-log
1 1 0.01 45 92
2 0.10 56 144
3 0.17 67 195
4 5 0.01 99 200
5 0.10 110 246
6 0.17 119 288
7 10 0.01 139 280
8 0.10 150 326
9 0.17 159 365
T O P I C 4
Health aspects / Pathogens and micro pollutants 482
Figure 2. Aquifer zones, based on piezometric gradient and well yield, for which well separation distances have been calculated
C O N C L U S I O N S
Conservative simplifying assumptions about solute transport and degradation in aquifers have been used to deter- mine distances over which a preconditioned minimum level of pathogen inactivation occurs in aquifers during ASR.
Maps of hydraulic gradient and well yield have been used to characterize regions in the shallow aquifer system of the Bandung Basin within which the spatial extent of pathogen inactivation is expected to be consistent. The results obtained indicate that 1-log removal distances vary from 45 m to 150 m between the zones, and for 4-log removals from 92 m to 326 m. This suggests that individual domestic scale ASR with water containing pathogens is quite likely to influence groundwater quality at nearby wells. Hence at domestic scale, roof runoff only should be ad- mitted to injection wells, to reduce the level of dependence on the aquifer for water treatment. The well separation required for 1-log removal is approximately half that required for 4-log removal. The smallest acceptable separation distances occur in flat low yielding areas that occur in the central part of the Bandung Basin. The more problematic areas, requiring larger separation distances between wells occur on the margins of the Bandung Basin where the aquifer is higher yielding and hydraulic gradients are steeper. Given these distances are larger than the average size of land holdings, any landholder injecting water of poor quality could adversely impact other groundwater users.
The techniques developed and utilised in this study may have widespread application in the design of ASR systems.
AC K N O W LE D G M E N T S
This work was supported by the Department of Education, Science and Training on a National Priority Reserve Project and funded by the Department of Education, Employment, Training and Youth Affairs. The project was administered by Carolyn Vicary of the Research Grants Section, Flinders University of South Australia. Assistance from Diah Wihardini (University of Adelaide) and Dr. Edi Utomo (LIPI) is sincerely appreciated.
R E F E R E N C E S
Bukit, Nana Terangna (1995). Water quality conservation for the Citarum River in West Java; Wat. Sci. Tech. 31, (9) 1 – 10.
Dillon, P. J., Miller, M., Fallowfield, H. and Hutson J. (2002). The potential of riverbank filtration for drinking water supplies in relation to microsystin removal in brackish aquifers. Journal of Hydrology, 266, (3-4) 209 – 221.
Dillon, P., Toze, S., Pavelic, P., Vanderzalm, J., Barry, K., Ying, G-G., Skjemstad, J., Nicholson, B., Miller, R., Correll, R., Prommer, H and Stuyfzand, P. (2005). Water quality improvements during aquifer storage and recovery at ten sites. (This volume).
Fildebrandt, S., Pavelic, P., Dillon, P. and Prawoto, N. (2003). Recharge enhancement using single or dual well sys- tems for improved groundwater management in the Bandung basin, Indonesia. CSIRO Technical Report 29/03, May 2003. http: //www.clw.csiro.au /publications /technical2003 /tr29-03.pdf.
Miller, R., Correll, R., Dillon, P. and Kookana, R. (2002). ASRRI: A predictive model of contaminant attenuation during aquifer storage and recovery. In Management of Aquifer Recharge for Sustainability, (ed. P J. Dillon), 69 – 74. A.A.Balkema.
Rhebergen, W. and Dillon, P. (1999). Riverbank filtration models for assessing viability of water quality improvement.
Centre for Groundwater Studies Report No. 90.
Soetrisno, S. (1998). Impacts of Urban and Industrial Development on Groundwater, Bandung, West Java, Indonesia, http://www.geocities.com/Eureka/Gold/1577/ paper_list_eng.html.
Suhari, S. and Siebenhuner, M. (1993). Environmental geology for land use and regional planning in the Bandung Basin, West Java, Indonesia; Journal of Southeast Asian Earth Sciences, 8, (1-4), 557 – 566, Pergamon Press, Great Britain.
Supriyo, A., Bambang, S. and Soetrisno, S. (1999). Aquifer storage and recovery for water conservation in Bandung Basin; 2nd CGS National Short Course on ASR, Adelaide, 27– 29 October 1999.
Suryantoro, Ir.S. (1999). Groundwater resources management in Indonesia; paper presented at the National Seminar on the decentralization of water resources management in Indonesia Sept.4th.
Toze S. (2004). Pathogen survival in groundwater during artificial recharge. In: Wastewater Re-use and Groundwater Quality. Proceedings of IUGG2003 Symposium HS04, Sapporo, July 2003. IAHS Publication No. 285, pp. 70– 84.
T O P I C 4
Health aspects / Pathogens and micro pollutants 484
Abstract
Cyanobacterial toxins are substances produced by cyanobacteria that occur in surface waters world wide. The most common group of cyanobacterial toxins is the group of structurally similar microcystins (MCYST).
Sand passage as used in slow sand filtration, artificial recharge and bank filtration has shown to be effective in eliminating microcystins in many cases. For secure drinking water production from surface waters infested by microcystins removal has to be ensured in a wide variety of cases met in the field.
It was therefore the aim of experiments in technical and semi-technical scale on the UBA’s experimental field in Berlin to test some worst case scenarios for the reliability of microcystin elimination during sand passage.
Experiments were conducted with virgin sand (no previous contact to MCYST) and high filtration rates as well as under anaerobic conditions. The results show that the greatest problem for MCYST elimination can be found under anaerobic conditions as degradation is not complete and may lead to harmful residual concentrations.
Keywords
Microcystins, underground passage, worst case scenarios, field scale experiments.
I N T R O D U C T I O N
Cyanobacterial toxins are substances with high acute and chronic toxicity produced by cyanobacteria or ‘blue- green-algae’ that can be observed in surface waters world wide. In case of cyanobacterial blooms that often develop in eutrophic water bodies, high concentrations of cyanobacterial toxins occur, with microcystins (MCYST) being the group of substances most frequently found. Under normal conditions over 90% of the toxins are contained within the cells. There have however been reports on high extra-cellular concentrations e.g. in case of aging popu- lations and /or sudden cell lysis.
Passage through biologically active strata, mostly sand, has shown to be effective in eliminating cyanobacterial cells as well as extracellular microcystins from surface water in many cases (Lahti and Hiisvirta, 1989; Sherman et al., 1995; Grützmacher et al., 2002). As aging populations accumulating on sediments may be assumed to show increased lysis and toxin release, an assessment of the reliability of the elimination of dissolved toxin for drinking water production is important. Elimination has to be secure under a variety of conditions met in the field. For this reason different worst case scenarios were simulated within the interdisciplinary NASRI (Natural and Artificial Systems for Recharge and Infiltration) research project dealing with river bank filtration processes. The aim was to identify the basic conditions under which sand and sediment passage can securely eliminate microcystins from sur- face water so that no further drinking water treatment has to take place.
The aim of a first series of experiments conducted in 2003 on a slow sand filter was to test whether a combination
elimination by sand passage?
G. Grützmacher, G. Wessel, I. Chorus and H. Bartel
of 3 worst-case conditions (SSF2: virgin sand that had no previous contact to MCYST, missing colmation layer and with 2.4 m /d high filtration velocities) leads to lower removal rates than experiments carried out with lower filtration velocities (SSF5: 1.2 m /d and SSF6 : 0.6 m /d) and clogging. The second series of experiments conducted in 2003 and 2004 on semi-technical scale enclosures were carried out under aerobic and anaerobic conditions and had the aim to show the influence of redox conditions on MCYST degradation.
M E T H O D S
E
Ex xp pe erriim me en ntta all m me etth ho od dss
• SSF experiments
The experiments were conducted on a technical scale slow sand filter (SSF) on the UBA’s (German Federal Environmental Agency) experimental field for simulation of underground passage (Bartel and Grützmacher, 2002).
The MCYST applied was extracted from a mass culture of Planktothrix agardhiiHUB 076 by centrifugation and freeze thawing in order to release the mainly cell-bound, highly water soluble microcystins. The freeze thawed extract was homogenized and then centrifuged to remove the cell debris and stored frozen.
In preparation of the experiments the flow rate through the SSF was adjusted to the desired amount, corresponding to filtration velocities of 2.4 m /d, 1.2 m /d and 0.6 m /d. The tracer and the MCYST were applied by spraying them evenly across the water reservoir with a hose from a barrel containing the concentrated substances diluted with 100 L of tap water. The tracer applied was sodium chloride (NaCl) so that the sampling intensity in the different sampling points could be addapted by observing the electrical conductivity (EC). Care was taken, not to raise the electrical conductivity by more than 10 % (Grützmacher et al., in press (a)). The resulting MCYST concentrations in the water reservoir amounted to 10 µg / L ± 2 µg /L.
• Enclosure experiments
Enclosure experiments were conducted in semi-technical scale columns for simulation of underground passage with an overlying water reservoir and a filter area of 1 m2 (see Grützmacher et al. in press (a) for details). Samples were taken from the water reservoir, from 40 cm and 80 cm depth and from the effluent (after 100 cm of sand passage).
One experiment (E5) was conducted under strictly aerobic conditions (oxygen was detected in the effluent at con- centrations of 11.7 mg / L ± 0.3 mg / L or 98% ± 2%). For the second experiment (E9) anaerobic conditions were induced by adding biodegradable DOC (acetic acid) continuously with a resulting concentration of 0.3 mmol / L additional DOC (resulting DOC concentration in water reservoir: 9.6 mg/L). After 4 days of continuous dosing oxygen could not be detected in 40 cm depth, after 9 days, no more oxygen was found in the effluent. After 3 weeks the redoxpotential (EH) had dropped to values < 0 mV in all 3 sampling points and iron and manganese reduction was observed, thus implying strictly anaerobic conditions after 40 cm of sand passage. Due to nitrate concentrations below detection limit in the inlet (< 0.1 mg / L ) denitrification not was observed.
The enclosure experiments were carried out similar to those on the slow sand filter, by applying a pulse of MCYST obtained from the mass culture of Planktothrix agardhii together with NaCl as a tracer to the water reservoir.
Samples were taken in regular intervals and analysed for MCYST by ELISA.
T O P I C 4
Health aspects / Pathogens and micro pollutants 486
A
An na allyyttiicca all m me etth ho od dss
Microcystin analysis was carried out by ELISA (Enzyme-Linked ImmunoSorbent Assay) and HPLC (High Per- formance Liquid Chromatography) with a photodiode array detector. Usually samples were first tested for their microcystin content using the ELISA and selected ones subsequently analyzed by HPLC to verify the results and distinguish the different microcystin variants.
• Sample preparation
The water samples for analysis of extracellular microcystins were filtered (RC 55, pore size 0.45µm) and stored deep frozen (–20°C). After thawing they were either analyzed directly (ELISA) or enriched by solid phase extraction (SPE) over C18-cartriges.
• ELISA
The ELISA used was a generic microcystin immunoassay based on monoclonal antibodies against the unusual Adda group characteristic for microcystins, nodularins and certain peptide fragments (Zeck et al., 2002). The measuring range lies between 0.1 µg/L and 1.0 µg/L. Samples from the water reservoir during the first phase of the experi- ments had to be diluted with deionized water. Each value was determined as the average of two parallels taken from the same sample.
• HPLC
After C18-SPE (see above) the microcystin variants were analysed by HPLC / p hotodiode array detection and iden- tified by means of their characteristic UV-spectra (Lawton et al., 1994). Depending on the water quality the detection limits ranged from > 1µg/L to 0.05 µg/L.
R E S U LT S A N D D I S C U S S I O N
S
SS SFF e ex xp pe erriim me en ntt w wiitth h h hiig gh h ffiillttrra attiio on n rra atte e a an nd d vviirrg giin n ssa an nd d
The relative concentrations of the main MCYST variant (demethylated MCYST-RR) applied during the SSF- experiment related to the initial maximum of 10.5 µg / L as well as the calculated dilution curve (confirmed by two tracer tests) are shown in Figure 1. This data confirms that this MCYST-variant is not degraded in the water reservoir during the mean residence time (half life of exponential input curve) of about 3 hours.
In total the MCYST recovered in the effluent amounted up to 24% of the total MCYST applied (356 mg). The curve shown in Figure 2 describes the modelled breakthrough (VCXTFIT, Nützmann et al., in press) with a retardation coefficient of 2.6 and a degradation rate of 0.17 h–1(which corresponds to a half life of 4,2 h).
Table 1 shows the retardation coefficients and degradation rates observed during all three SSF experiments.
Surprisingly experiment SSF2 with the highest filtration rate, virgin sand and virtually no clogging shows the high- est retardation coefficient and degradation rate. This might be due to reactions (e.g. irreversible sorption) that take place only during first contact with MCYST.
Table 1. Retardation coefficients (R) and degradation rates (λ) obtained by modelling slow sand filter experiments with different filtration velocities and clogging situations
Experiment #
Filtration velocity
[m/d] R
λ
[h–1] Remarks
SSF2 2.4 2.6 0.17 virgin sand, no clogging
SSF5 1.2 1.3 0.04 some clogging
SSF6 0.6 1.5 0.05 clogged
T O P I C 4
Health aspects / Pathogens and micro pollutants 488
0,0 0,1 0,2 0,3 0,4 0,5 0,6 0,7 0,8 0,9 1,0
0 10 20 30 40 50
time after application (h) c/c0
MCYST-RR variant dilution curve
Figure 1. Concentration of demethylated MCYST-RR (main MCYST variant of mass culture) in the water reservoir during the slow sand filter experiment
0,00 0,05 0,10 0,15 0,20 0,25 0,30
0 10 20 30 40 50
time after application (h) sum recovered MCYST (c/c0)
measured values model
Figure 2. Sum of recovered MCYST (all variants) in the effluent during the SSF-experiment (model: R = 2,6; λ= 0,17 h–1)
E
En nccllo ossu urre e e ex xp pe erriim me en ntt u un nd de err a ae erro ob biicc a an nd d a an na ae erro ob biicc cco on nd diittiio on nss
The maximum MCYST concentrations measured by ELISA in the four sampling points during the aerobic (E5) and the anaerobic (E9) experiment in relation to the maximum input concentration (E5: 28.6 µg/L and E9: 10.3 µg/L) are shown in Figure 3. Due to clogging during the anaerobic experiment, the residence times for 80 cm and the effluent were much higher than during the aerobic experiment. However MCYST reduction in the anaerobic exper- iment reached only 60% maximum in 80 cm depth with no further reduction in the following 20 cm of filter material. The degradation rates calculated by fitting exponential decay curves to the measured values amount to 0.235 h– 1(aerobic) and 0.22 h– 1(anaerobic), thus showing no substantial difference. The main difference is the residual of 3.8 µg/L (35%) under anaerobic conditions that does not seem to be degraded any further after 12 h contact time.
Indications of reduced biodegradation of MCYST und anaerobic conditions have so far been reported largely from laboratory experiments (Grützmacher et al., in press (b); Holst et al., 2002). It was therefore important to verify this in nearly natural settings and to show that in spite of an aerobic zone (which is present in most cases where surface water infiltrates into the aquifer), MCYST is not degraded completely, thus leading to a residual con- centration that remains in the water.
C O N C L U S I O N S
Two worst case scenarios for MCYST elimination during underground passage were tested in technical and semi- technical scale experiments. The first scenario with conditions assumed to be quite unfavourable for degradation, i.e. virgin sand and high flow velocity, did not show reduced elimination of MCYST compared with the other scenarios tested (preconditioned sand, lower flow velocities). In contrast, under the presumably unfavorable condi- tions, higher retardation and degradation was observed. For the setting simulated (aerobic sand passage) only a few days (< 5) are sufficient to reduce maximum extracellular MCYST concentrations of 100 µg/L to values lower than the provisional WHO guideline value of 1 µg / L MCYST-LR (WHO 1998).
Figure 3. Maximum MCYST concentrations (measured by ELISA) during the aerobic and anaerobic enclosure experiment
The second scenario tested was the simulation of anaerobic conditions that are likely to be met in the field during and after a cyanobacterial bloom, due to high organic carbon concentrations in the surface water. After an initial reduction of about 65% of the maximum input concentration no further degradation was observed under anaerobic conditions between 12 and 25 h contact time. Under aerobic conditions, in contrast, MCYST was reduced to values near the detection limit after 12 h contact time. Holst et al. (2003) carried out laboratory batch experiments on MCYST degradation under anaerobic conditions with addition of nitrate and glucose to stimulate denitrifying bacteria. The results showed that MCYST was not degraded in the sterile experiment and only little degradation was observed in the series without additional nutrients (to about 70% of the initial value within one day and then no further decline). In the experiments with additional nutrients, MCYST was degraded to less than 20 % of the initial value within one day. This implies that under anoxic conditions MCYST degradation is limited to denitrifying, high nutrient conditions. Thus in our experiments the lack of nitrate might have been the reason for incomplete MCYST degradation as concentrations measured in the water reservoir were already below the detection limit.
AC K N O W LE D G E M E N T S
We thank the Berliner Wasserbetriebe (BWB) and Veolia Water for sponsoring the NASRI research project during which the experiments were conducted. We would also like to thank I. Flieger for the HPLC analyses as well as H.W. Althoff, T. Starzetz and T. Köhler for their help during the technical scale experiments.
R E F E R E N C E S
Bartel H. and Grützmacher G. (2002). Elimination of microcystins by slow sand filtration at the UBA’s experimental field. In: Riverbank filtration: Understanding Contaminant Biogeochemistry and Pathogen Removal, C. Ray (ed.), Kluwer Academic Publishers, pp. 123 – 133.
Grützmacher G., Bartel H. and Wiese B. (in press (a)). Simulating Bank Filtration and Artificial Recharge on a Tech- nical scale. Conference proceedings ISMAR 2005.
Grützmacher G., Wessel G., Chorus I., Bartel H. and Holzbecher E. (in press (b)). On the behaviour of microcystins in saturated porous medium. Conference proceedings ISMAR 2005.
Grützmacher G., Böttcher G., Chorus I. and Bartel H. (2002). Removal of microcystins by slow sand filtration.
Environmental Toxicology, 17(4), 386 – 394.
Holst T., Jørgensen N.O.G., Jørgensen C. and Johansen A. (2002). Degradation of microcystin in sediments at oxic and anoxic, denitrifying conditions. Wat. Res., 37, 4748 – 4760.
Lawton L.A., Edwards C., Codd G.A. (1994). Extraction and high-performance liquid chromatographic method for the determination of microcystins in raw and treated waters. Analyst, 119, 1525 – 1530.
Lahti K., and Hiisvirta L. (1989). Removal of cyanobacterial toxins in water treatment processes: Review of studies conducted in Finland. Water Supply, 7, 149 – 154.
Nützmann G., Holzbecher E., Strahl G., Wiese B., Licht E., Grützmacher G. (subm.). Visual CXTFIT a user-friend- ly simulation tool for modelling one-dimensional transport, sorption and degradation processes during bank fil- tration. Conference proceedings ISMAR 2005.
Sherman P., Tully I. and Gibson H. (1995). Removal of cyanobacterial cells and toxins from drinking water with bio- logically active filters. Proceedings of the 16th Federal AWWA Convention, April 2-6, 1995, Sydney, Australia, 587 – 592.
World Health Organization (1998). Guidelines for Drinking Water Quality, 2nd edition, Addendum to Volume 2, Health Criteria and other supporting Information, WHO, Geneva.
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Health aspects / Pathogens and micro pollutants 490
Abstract
Microcystins (MCYST) are a group of toxic substances produced by cyanobacteria (‘blue-green-algae’). In case of cyanobacterial blooms microcystin concentrations in surface waters may reach values far above the value proposed as provisional guideline for drinking water by the WHO of 1 µg/L for MCYST-LR. For drinking water production via underground passage it is therefore necessary to ensure removal to a large extent.
For this reason experiments with extracellular microcystins were conducted in the laboratory as well as in a natural setting on the UBA’s (German Federal Environmental Agency) experimental field for simulation of under- ground passage.
Laboratory batch experiments showed that adsorption of microcystins can be neglected in sandy material (kd< 1 cm³/g). Batch and column experiments identified biodegradation as the predominant elimination process in these sediments. The degradation rates derived from laboratory column experiments as well as semi-technical scale enclosure experiments varied between 0.2 d–1and 18 d–1. In the worst case this means a half life of 2.8 days, so that under aerobic conditions contact times of several days should be sufficient to eliminate MCYST to an extent safe for use as drinking water.
Keywords
Degradation rates, field scale experiments, kd-values, laboratory experiments, microcystins, underground passage.
I N T R O D U C T I O N
Microcystins (MCYST) are a group of structurally similar toxic substances produced by cyanobacteria or ‘blue- green-algae’. They occur frequently in cyanobacterial blooms which can be observed in surface waters world wide (Bartram et al. 1999). When using these surface waters for drinking water production via bank filtration or artificial recharge elimination of microcystins by underground passage has to be ensured as the WHO has proposed a pro- visional guideline value for drinking water of 1 µg/L for MCYST-LR (WHO 1998).
Microcystins are subject to different processes in the porous medium, such as degradation, sorption and desorption.
The relevance of the different processes depends on various conditions: on the properties of the porous medium (grain size distribution, organic content), as well as on the redox state (aerobic / anoxic /anaerobic), the pH, temperature, etc..
Within different interdisciplinary research projects dealing with river bank filtration processes several experiments were conducted. The aim of the paper is to present the outcome of the different measurements and to compare the main sorption and degradation parameters.
G. Grützmacher, G. Wessel, H. Bartel,
I. Chorus and E. Holzbecher
M E T H O D S
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Ex xp pe erriim me en ntta all m me etth ho od dss
Experiments were carried out in laboratory batch- and column systems (Grützmacher et al. in prep.), as well as in technical scale enclosures and slow sand filters under different clogging conditions (Grützmacher et al. in press).
Enclosures are large scale columns (diameter: 1.13 m) embedded in a slow sand filter with a water reservoir of 40 cm and a sand filter of 1 m depth. This arrangement was chosen in order to expose the column to the most close-to-real conditions, the only difference being the smaller size of the filter bed. Extra-cellular microcystins were applied in realistic concentrations of about 1 to 10 µg / L . The microcystins were either prepared from commercially available standards (MCYST-LR, -YR and -RR) or obtained by freeze-thawing a concentrate (aqueous extract) of a mass culture of Planktothrix agardhiiHUB 076 that is currently run by the UBA in Berlin. The con- centrate was subsequently centrifuged and the supernatant deep frozen for further conservation. The microcystin concentrations in the undiluted extract reached 50 mg/L, the DOC amounted up to 4 g / L . For the experiments the extract was diluted about 1 : 5,000. MCYST concentrations in the samples from the laboratory experiments were usually determined by ELISA, those taken during enclosure experiments by HPLC (see below).
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Microcystin analyses were carried out by ELISA (Enzyme-Linked ImmunoSorbent Assay) and HPLC (High Performance Liquid Chromatography) with a photodiode array detector. Usually samples were first tested on their microcystin content using the ELISA and selected ones subsequently analyzed by HPLC to verify the results and distinguish the different microcystin variants.
• Sample preparation
The water samples for analysis of total microcystins were deep frozen and thawed in order to release cell-bound microcystins, filtered by membrane filters (RC 55, pore size 0.45 µm) and either analyzed directly (ELISA) or enriched by solid phase extraction (SPE) over C18-cartriges and additional silica clean-up according to Tsuji et al.
(1994). For analysing the dissolved microcystins samples were filtered directly after sampling and then deep frozen for subsequent analysis by ELISA or HPLC.
• ELISA
The ELISA used was the EnviroGard Plate Kit (offered by Coring System Diagnostix GmbH, Germany) with a measuring range between 0.1 µg/L and 1.6 µg / L . Samples from the water reservoir during the first phase of the experiments had to be diluted with deionized water. Each value was determined as the average of two parallels taken from the same sample. The preparation of the plates was done according to the instructions by the producer.
• HPLC
After C18-SPE (see above) and additional silica cleanup the microcystin variants were analysed by HPLC / photo- diode array detection and identified by means of their characteristic UV-spectra (Lawton et al. 1994). Depending on the water quality the detection limits ranged from 0.05 µg / L to > 1 µg / L .
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Parameter estimation for tracers was performed using Visual CXTFIT (Nützmann et al. in press), a graphical user interface for the CXTFIT code, developed by Toride et al. (1995). Retardations and degradation rates were
T O P I C 4
Health aspects / Pathogens and micro pollutants 492
estimated using basic MATLAB (2002) package in connection with the optimization toolbox. The numerical pdepe-solver, included in MATLAB, was used for direct modelling within the estimation procedure.
R E S U LT S
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Ad dsso orrp pttiio on n iisso otth he errm mss
Batch experiments for determination of adsorption parameters were carried out with different sediments (material from a slow sand filter, aquifer sediment as well as pure quartz sand) and MCYST-variants (MCYST-LR, -YR and -RR). The resulting kd-values are presented in Table 1. Although slight differences in adsorption can be observed depending on the MCYST-variant (MCYST-RR adsorbing more readily than MCYST-YR and -LR), the most impor- tant parameter seems to be the texture of the sediment as the aquifer sediment contains a relatively high portion of organic substance as well as clay and silt.
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De eg grra ad da attiio on n e ex xp pe erriim me en nttss
Parallel sterile (autoclaved) and non sterile batch experiments carried out with natural filter sand (SSF-material) and surface water showed virtually no decrease of MCYST-concentration in the sterile experiment. Biodegradation is therefore the most important elimination process for MCYST in contact with sandy sediment (Figure 1).
For further quantification of degradation rates closed loop column experiments were conducted with the same material (SSF-material) and different MCYST-variants under aerobic conditions (Figure 2). The results showed similar degradation rates between 1.33 d–1and 1.74 d–1(the half lives correspondingly amount to around 10 h).
In order to test MCYST degradation under field conditions an aqueous extract from the mass culture of Planktothrix agardhii was applied to the water reservoir of an enclosure on the UBA’s experimental field. Here sediment passage can be simulated in a semi-technical scale (filter area: 1 m², filter depth: 1 m). Samples were taken from the water reservoir, from 20 cm, 40 cm, 60 cm, 80 cm depth and from the effluent after 100 cm sand passage.
Simultaneously applied tracer (NaCl) gave values for pore velocity and dispersion coefficients (Grützmacher et al.
in press). On the basis of these values retardation factors and degradation rates were calculated, that took the measured input concentrations of MCYST into account. An example for a modelled curve as well as the measured input and output values is given in Figure 3.
Table 1. kd-values determined in batch experiments and important sediment properties (LOI (loss on ignition), silt and clay-content)
Sediment LOI (%) Silt and clay (%) MCYST-variant kd-value (cm³/g)
Slow sand filter material 0.2 < 0.1
MCYST-LR 0.25
MCYST-YR 0.22
MCYST-RR 0.85
Aquifer sediment 0.8 max. 3 MCYST-LR 11.6
Quartz sand < 0.05 0 MCYST-LR 0.08
Two enclosure experiments were conducted on the same enclosure with an interval of one month, allowing some clogging to take place in the meantime. The hydraulic conductivity decreased about 11% during this time (from 1.8 x 10–5m/s to 1.6 x 10–5m/s, Grützmacher et al. in press). The results of the modelling of the two experiments with MCYST are given in Table 2. The highest retardation and degradation rates were observed in the uppermost centimetres in both experiments. Both retardation and degradation tend to decrease with depth. There is, however, a zone of lower retardation (E3) and lower degradation (E2) between 20 cm and 40 cm. This might be due to vary- ing saturation, as unsaturated conditions were observed with proceeding clogging. The last zone between 80 cm and the outlet also shows surprisingly high retardation coefficients. This may be due to column-end-effects arising from the gravel drainage layer.
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Health aspects / Pathogens and micro pollutants 494
Figure 1. Batch experiments with MCYST (in surface water with filter sand) under sterile and non-sterile conditions (average and rage of 2 parallels)
Figure 2. Results of non-sterile closed loop column experiments with different MCYST-variants (degradation rates in brackets)
In order to compare adsorption parameters obtained from laboratory batch experiments to those calculated on the basis of field enclosure experiments kd-values were calculated on the basis of the retardation coefficients according to (1), with neranging from 0.16 to 0.4 and ρsamounting to an average of 1.65 g/cm³.
Eqn. (1) The resulting average kd-value for experiments E2 and E3 amounted to 0,14 cm3/ g and 0.07 cm3/ g , respectively.
This is less than 50% of the values obtained in the batch experiments. This might be due to the different ratio of water to sediment (batch experiment 1:1, enclosure experiment about 3:1).
(
1)
* −
⎟⎠
⎜ ⎞
⎝
= ⎛n R k
s e
d ρ
0 5 10 15 20 25 300
1 2 3 4 5 6
Mcyst concentration [g/l] at inlet
0 5 10 15 20 25 30
0 0.5 1 1.5 2 2.5
Time t [h]
Mcyst concentration [g/l] at outlet
measured modelled
Figure 3. Measured and modelled MCYST-concentrations in 20 cm depth during enclosure experiment E2 (diamonds: MCYST-concentration in the inlet);
parameters for the given situation are: ν= 0.32 m / h , αL = 0 . 0 0 6 3 m , R = 1 . 9 7 , λ = 0 . 74 l / h Table 2. Retardation coefficients and degradation rates calculated for extracellular MCYST
during two enclosure experiments E2 and E3
Experiment
retardation coefficient (R)
degradation rate (λ) [1/d]
E2 E3 E2 E3
water reservoir -> 20 cm 1.97 1.76 17.8 13.0
20 cm -> 40 cm 1.9 1.25 5.76 7.44
40 cm -> 60 cm 1.0 1.36 18 4.56
60 cm -> 80 cm 1.0 1.14 13.0 0.24
80 cm -> outlet 1.58 1.3 1.68 1.68
The degradation rates observed during the laboratory experiments also differ from those calculated from the results of the enclosure experiments. They are, however, in the same order of magnitude and in some depths of the enclo- sure (e.g. 80 cm to outlet) they correspond well.
C O N C L U S I O N S
Under aerobic conditions biological degradation during sand passage was shown to be effective in eliminating extra- cellular microcystins. Contact times of only a few hours seem to be sufficient to reduce concentrations typical for cyanobacterial blooms in the water body to values below the WHO guideline value. Similar degradation rates were determined by Christoffersen et al. (2002), however for surface water without sediment contact. Batch experiments conducted by Holst et al. (2001) showed much slower degradation (half lives around 25 d). These experiments were however carried out with aquifer material without previous contact to MCYST and at very low temperatures (4°C).
Adsorption is only relevant in material with clay or silt content, which is supported by the findings of Miller (2000). If adsorption takes place it can lead to longer contact times which supports elimination by biological degra- dation. There are however indications of reduced degradation rates under anoxic and anaerobic conditions. These conditions are subject to further investigations that are currently being carried out.
Laboratory experiments can give an impression of which processes take place under certain, controlled conditions.
However kd-values and degradation rates calculated on the basis of laboratory experiments only have to be treated with caution. Before transferring these to the field it is suggested to conduct field observations or technical scale experiments for further verification.
AC K N O W LE D G E M E N T S
We thank the Germen Ministry for Research and Education (BMBF) for funding the laboratory experiments (Grand No. 02WT9852 / 7 ) and the Berliner Wasser Betriebe (BWB) / Veolia Water for sponsoring the project NASRI, in the course of which the enclosure experiments were carried out. Special thanks to I. Flieger for her diligent analytical work as well as H.W. Althoff, T. Starzetz and T. Köhler for their help during the field scale experiments.
R E F E R E N C E S
Bartram, J., Carmichael, W.W., Chorus, I., Jones, G. and Skulberg, O.M. (1999). Introduction to ‘Toxic Cyanobacteria in Water’. In: Toxic Cyanobacteria in Water, I. Chorus and J. Bartram (eds.), F&FN Spon, London, pp. 1 – 13.
Christoffersen K., Lyck S. and Winding A. (2002). Microbial activity and bacterial community structure during degradation of microcystins. Aq. Microb. Ecology, 27: 125 – 136.
Grützmacher, G., Bartel, H. and Wiese B. (in press). Simulating Bank Filtration and Artificial Recharge on a Technical Scale. – Proceedings of the 5thInternational Symposium on Management of Aquifer Recharge.
Grützmacher, G., Wessel, G. and Chorus, I. (in prep.): Laboratory Experiments on Microcystin Elimination during Sediment contact.
Holst T, Jergensen NOG, Jorgensen C, Johansen A (2003) Degradation of microcystin in sediments at oxic and anoxic, denitrifying conditions. Wat. Res.37(19), 4748 – 4760.
Lawton L.A., Edwards C., Codd G.A. (1994). Extraction and high-performance liquid chromatographic method for the determination of microcystins in raw and treated waters. Analyst, 119, 1525 – 1530.
MATLAB (2002, Release 13), The MathWorks, Inc., 3 Apple Hill Drive, Natrick, MA 01760-2098, USA
T O P I C 4
Health aspects / Pathogens and micro pollutants 496
Miller M.J. (2000). Investigation of the removal of cyanobacterial hepatotoxins from water by river bank filtration. PhD Thesis, Flinders University, Adelaide, Australia.
Nützmann, G., Holzbecher, E., Strahl, G., Wiese, B., Licht, E., Grützmacher, G. (in press). Visual CXTFIT a user- friendly simulation tool for modelling one-dimensional transport, sorption and degradation processes during bank filtration. – Proceedings of the 5th International Symposium on Management of Aquifer Recharge.
Toride N., Leij F.J., and van Genuchten M. Th. (1995). The CXTFIT code for estimating transport parameters from laboratory or field tracer experiments, U.S. Salinity Lab., Agric. Res. Service, US Dep. of Agric., Research Report No. 137,Riverside (CA).
World Health Organization (1998). Guidelines for Drinking Water Quality, 2nd edition, Addendum to Volume 2, Health Criteria and other supporting Information, WHO, Geneva.
Abstract
The UBA’s experimental field on the outskirts of Berlin offers a unique possibility of simulating bank filtration, artificial recharge and slow sand filtration on a technical scale.
The site consists of a storage reservoir (pond) with an adjacent artificial aquifer consisting of sand and gravel.
Additionally the surface water can be conducted into 4 infiltration basins (two slow sand filters and two aquifer infiltration ponds). Three enclosures as well as large scale columns can be used for shorter and longer term simu- lation of groundwater transport. The whole site is separated from the surrounding aquifer by a layer of clay. A variety of physico-chemical parameters can be measured continuously and observed online.
The travel times for the bank filtration passage determined by tracer experiments range from a few days to a maximum of 3 weeks. In the enclosures, infiltration ponds and large scale columns contact time can be varied between a few hours up to 3 months.
Keywords
Artificial recharge, bank filtration, experiments, slow sand filtration, technical scale.
I N T R O D U C T I O N
The UBA (Umweltbundesamt or Federal Environmental Agency of Germany) is running an experimental field on the outskirts of Berlin. Part of the site consists of a storage pond with adjacent artificial aquifer and infiltration ponds that offer a unique possibility of simulating bank filtration, artificial recharge and slow sand filtration on a technical scale.
In order to characterize groundwater flow in the system many tracer tests were conducted revealing the hydraulic properties of the facilities as well as temporal variations in hydraulic conductivity due to clogging.
M E T H O D S
E
Ex xp pe erriim me en ntta all ffa acciilliitte ess
The site consists of a storage reservoir (pond) with a water volume of about 3,500 m3and an adjacent artificial aquifer consisting of fine and coarse gravel (Figures 1 and 2). Additionally the surface water of the pond can be conducted into four infiltration basins each with a square basal surface of 72 m2. Two of these basins are sealed by concrete at the bottom with a sand depth of 0.8 m (here water is collected by 3 parallel drainages at the bottom), two have direct contact to the underlying sediments and can be operated as aquifer infiltration ponds. For smaller scale experiments three enclosures with a surface area of 1 m2and medium grained sand filling are installed in one of the infiltration basins (Figure 3). The whole site is separated from the surrounding aquifer by a layer of clay thus
Simulating bank filtration and artificial recharge
on a technical scale
Gesche Grützmacher, Hartmut Bartel and Bernd Wiese
VV
forming an independently operable hydraulic system so that experiments with toxic substances can be carried out without adverse effects on the environment.
A variety of physico-chemical parameters (e.g. pH, temperature, oxygen, redoxpotential, electrical conductivity (EC), TOC, total bound nitrogen (TNb) as well as fluorescence for determination of algal biomass) can be meas- ured continuously and observed online.
The sand filling of the slow sand filter and the enclosures was replaced in autumn 2002. The kf- value calculated on the basis of the grain size distribution according to Beyer (1964) is 7 * 10– 4m/s for the middle sand inside the slow sand filters and enclosures and 1 *10– 1m/s for the gravel drainage layer. In March 2003 the basins were flooded again and experiments were conducted until November 2003 (Table 1). In the slow sand filter the clogging layer was not removed throughout 2003 thus giving experimental results for growing clogging conditions. In April 2003, however, between experiment SSF1 and SSF3 bacterial slide holders were installed in one part of the filter with substantial disturbance of the filter sand up to a depth of about 30 cm.
15 observation wells
30.5 m
storage pond
> 2 m 1.5 - 2 m 1 – 1.5 m
55.4 m
45.5 m
88.3 m
slow sand filters (SSF)
& infiltration ponds
bank filtration
inlet SSF
SSF
17.5 m
30.5 m
> 2 m 1.5 - 2 m 1 – 1.5 m
55.4 m
45.5 m
88.3 m
bank filtration
SSF SSF SSF
17.5 m17.5 m
Figure 1. Map of the storage pond system (modified from Bartel & Grützmacher 2002)
5 m
2 m 7 m
22 m
observation wells
storage pond slow sand filter
about 4 m
gravel (8 / 20 mm) gravel (2 / 8 mm)
sand (0.8 / 2 mm)
gravel (32 / 56 mm) with drainage pipe concrete 5 m
2 m 7 m
22 m 2 m 5 m 7 m
22 m about 4 m
Figure 2. Cross-section of the artificial aquifer with the overlying storage pond
The hydraulic conductivity (kf) of the SSF was calculated according to (1) with vf(filtration velocity = average flow rate during the experiment divided by the average filter area of 60 m2) and i (gradient = surface water reservoir minus height of outlet divided by 0.8 m filter depth).
kf = vf / i (1)
The flow in the enclosures had to be interrupted several times due to technical problems. During the experiments hydraulic conductivity was determined according to (1) with an average filter area of 1 m2and i = p (bar) * 10.197 (m/bar) (p measured in front of the pump at the effluent). Pumping at Enclosure III was commenced shortly before the first experiment in July 2003 after a stagnant period of about 3 months. Enclosure II was used only in November 2003 after having remained flooded with interrupted flow for 8 months. After experiment No. E4 the upper layer of 0.5 cm of algae and debris was removed in order to observe the effect of this layer on substance removal.
TTrra acce err tte essttss
Parallel to each experiment simulating substance removal during underground passage on the artificial aquifer sys- tem tracer tests were conducted for characterization of groundwater flow. Usually chloride was used as a tracer by adding sodium chloride to the infiltrating water and measuring the electrical conductivity in the different sampling ports and effluents. Care was taken to raise the electrical conductivity not more than 10% above the background level (950 µS /cm in average) to avoid induction of hydrochemical reactions. The hydrochemical analyses showed this was successful with exception to a slight cation exchange reaction during tracer passage.
The tracer was added to the water reservoir in a pulsed application and pumps were installed in the water body throughout the experiment in order to reach a homogenous distribution. Subsequently inflowing water diluted the maximum tracer concentration in a rate depending on the adjusted flow rate thus giving an exponentially decreas- ing input concentration. The fitted first order exponential function was used as input for modelling using the pro-
T O P I C 4
Health aspects / Pathogens and micro pollutants 500
Table 1. Summary of parameters for the tracer tests in 2003
Experiment No. Location Date Filtration velocity [m/d]
SSF1 Slow sand filter 19.03.03 2.4
SSF3 Slow sand filter 23.04.03 2.58
SSF5 Slow sand filter 17.06.03 1.43
SSF6 Slow sand filter 19.11.03 0.59
E2 Enclosure III 05.08.03 1.25
E3 Enclosure III 09.09.03 1.12
E4 Enclosure II 11.11.03 1.09
E5 Enclosure II 25.11.03 1.22
Figure 3. Schematical cross-section of an enclosure
gram Visual CXTFIT (Nützmann et al., in press). By inverse modelling the values of dispersion and pore velocity could be determined serving as a basis for modelling the behaviour of reactive substances added. The sampling ports in the enclosures and slow sand filters were connected to peristaltic pumps and run continuously with low flow rates (60 to 600 ml / h).
The artificial aquifer as a whole was subject to one tracer experiment using bromide and a Gd-DTPA complex (Wiese et al., submitted). The tracers were applied to the storage pond by pumping a highly concentrated solution through a perforated hose that was arranged parallel to the aquifer’s bank. A current induced by air injection through a perforated hose next to the tracer hose provided mixing.
Samples from the 15 observation wells were taken with a peristaltic pump, ensuring the water in the tube was exchanged at least once. The data was modelled with Modflow-MT3DMS to obtain hydraulic properties of the 3-D structure and Visual CXTFIT to obtain transport properties of each borehole.
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Electrical conductivity was determined using a WTW conductometer (measuring range 1 to 100,000 µS /cm, ± 1 %).
Bromide concentrations were measured with an Ion chromatograph DX 500 (Dionex Coop.) according to DIN EN ISO 10304-1/2. Gadolinium was analysed with a method described by Bau and Dulski (1996).
R E S U LT S A N D D I S C U S S I O N
TTrra acce err tte essttss o on n ssllo ow w ssa an nd d ffiilltte errss a an nd d e en nccllo ossu urre ess
Figure 4 gives an example of the conductivities measured in the effluent as well as the curves modelled by using Visual CXTFIT. Table 1 lists the modelled pore velocities, dispersion lengths (αL) as well as the filter resistance for each of the experiments conducted on the slow sand filters and enclosures.
The slow sand filter experiments were conducted with three different filtration velocities (about 2.5 m/d, 1.4 m/d and 0.6 m/d). Therefore pore velocities changed correspondingly. Porosities are between 0.357 and 0.385. The
15 observation wells
30.5 m
storage pond
> 2 m 1.5 - 2 m 1 – 1.5 m
55.4 m
45.5 m
88.3 m
slow sand filters (SSF)
& infiltration ponds
bank filtration
inlet SSF
SSF
17.5 m
30.5 m
> 2 m 1.5 - 2 m 1 – 1.5 m
55.4 m
45.5 m
88.3 m
bank filtration
SSF SSF SSF
17.5 m17.5 m
Figure 4. Measured electrical conductivity and modelled output function for tracer experiment No. E2 (vf = 1.2 m/d)
