Figure 5. Japanese Ambient Air Cadmium Concentration and NiCd Battery Production
Table IV. Metal Emissions in Production of NiCd Electric Vehicle Batteries*
Grams Metal Emissions per KW-Hour Emission Sink Nickel Cadmium Cobalt
Air Water Land Total
3.15 2.28 Negligible 5.43
1.78 1.31 Negligible 3.09
0.08 0.05 Negligible 0.13
*Source: Geomet Technologies
Typical industrial NiCd batteries utilized for electric vehicle applications have an energy density of 50 Watt-Hours per kilogram which corresponds to weight per unit energy of 20 kilograms per Kilowatt-Hour. Thus the metal emission levels associated with the manufacture of industrial NiCd batteries are roughly 0.04% of the total weight of the battery, in reasonably close agreement with most present day estimates which place total metal emissions during battery manufacturing between 0.01% and 0.1%. The data sets in Tables III and IV agree reasonably well, particularly if the decreasing levels of metal emissions with time are considered.
Finally, a life cycle analysis conducted by SAFT (Cornu and Eloy 1995) on nickel-cadmium batteries for electric vehicle applications has established results comparable to those cited above. This study indicates that losses during the manufacturing process are likely to be on the order of 0.037% of battery weight for nickel and 0.008 - 0.019% of battery weight for cadmium, depending on the recycling options adopted. Once again, these estimates are consistent with other studies, and indicate very low environmental and human health impacts from the manufacturing stage of a battery's total life cycle analysis.
Thus, the emissions associated with the manufacture of battery systems, like those associated with the production of the primary raw materials, are generally quite low, probably less than 1% of the total potential emissions if the spent battery were discarded entirely into the environment after use. While most of the data presented above are relevant mainly to nickel-cadmium batteries, which have been heavily studied because of regulatory and environmental controversy, the same general conclusions apply to other battery systems in general with some variations. Primary raw material production and battery manufacturing , in general, contribute only a small fraction of the environmental or human health impact that might be encountered in unconsidered waste disposal.
Rechargeable batteries are long-lived products which, in general, may be used many times over if they are charged and discharged properly. Non-rechargeable batteries are shorter lived products, but in some cases have higher initial energy density than rechargeable batteries. Since various battery chemistries, in general, will have different operating voltages, energy ratings and cycle lives (if they are rechargeable), each individual battery system will have different total lifetime energy characteristics. Even within the same battery chemistry family, there will be variations to suit specific applications. Thus, an AA-sized NiCd battery may exhibit energy ratings from 500 milliampere-hours to 1,500 milliampere-hours depending on its intended use. Correspondingly, other properties, such as cycle life, may vary as well. In addition, total battery energy varies with battery size, the larger the battery in general the larger its total lifetime energy, other factors being equal. Therefore, it may be very difficult to establish an average set of performance characteristics for a battery family, but only to establish them for a very specific battery chemistry, size and type.
Battery performance is importance because in determining any human health or environmental impacts of battery systems, these must be normalized to a unit energy basis, as previously noted for emissions associated with battery manufacturing. Thus, any emissions during any stage of the life cycle of the battery system must be divided by the total lifetime energy of the battery to obtain results which allow comparison amongst battery systems. The total lifetime energy of a particular battery system is the product of its voltage, capacity and cycle life. Strictly speaking, charging efficiency and self-discharge characteristics should also be taken into account, but in most life cycle analyses to date, they have not been. For example, the basic performance parameters of an AA-sized NiCd battery are summarized in Table V.
Table V. Basic Performance Parameters of AA NiCd Battery
Range of Values
Voltage Capacity Total Energy Cycle Life (80%DOD) Total Lifetime Energy
1.2 Volts 0.5 to 1.2 ampere-hours 0.6 to 1.4 watt-hours 700 to 1200 cycles 420 to 1680 watt-hours
Nickel-cadmium and nickel metal hydride batteries both operate at 1.2 volts, whereas alkaline manganese batteries produce 1.5 volts and lead acid batteries 2.0 volts. Lithium-ion batteries have an unusually high voltage, above 3.0 volts, which gives them a high energy density. Thus, all three of the parameters mentioned above - voltage, battery capacity, and cycle life - will be instrumental in establishing the life cycle performance of a battery system. It is not just the composition of the battery alone which is important, and, as will be subsequently shown, it is the waste disposal option chosen for the battery which is perhaps even more important than either of these first two characteristics in determining life cycle impact.
During the normal use and maintenance of a battery system, they are neither destroyed nor dissipated nor do they emit any harmful substances. Battery systems may be sealed or vented. If they are sealed, then no emissions occur during normal use and maintenance. If they are vented, then water vapor, hydrogen gas or oxygen gas may be vented, depending on the system and whether it is charging or discharging. A 1994 report (Stockholm Environmental Institute 1994), for example estimated that the dissipation rates for both industrial and consumer NiCd batteries were 0.01 percent per year. The International Cadmium Association believes, based on surveys of its NiCd battery producer members, that the dissipation rates are virtually zero, or so low as to be undetectable.
However, a further consideration is the potential life cycle effect of each recharging cycle for the battery. The energy necessary to recharge a battery is generated by the primary power grid which generally operates on some form of fossil fuel. Combustion of fossil fuels result in the generation of greenhouse gases which can have an effect on a complete life cycle analysis, particularly if dirty fossil fuels are used or air pollution emission control devices are inadequate. In general, the emissions and life cycle effects associated with recharging are again small compared to those of battery disposal. One analysis (Schuckert et al. 1997) has measured the primary energy consumption during the production and utilization of both lead acid and nickel-cadmium batteries and their consequent effect upon carbon dioxide emissions and nitrous oxide emissions. In these cases, the amounts of energy required and greenhouse gases generated over the battery system's entire lifetime are lower for NiCd batteries than for lead acid batteries because of their higher cycle life, energy density and total lifetime energy even though the initial energy required to produce the NiCd battery is higher than to produce the lead acid battery.
In a life cycle impact analysis of battery systems, regardless of composition, performance and whether or not they are rechargeable, it is clearly the final disposal of the battery which determines its major environmental and human health impact. The emissions associated with all the stages up to the disposal of the battery are perhaps only 1% to 2% of the total potential emissions if the battery is simply discarded into the environment. This figures changes, of course, if the battery is disposed of in a controlled manner such that emissions are minimized. Nonetheless, disposal is the key step in determining total environmental or human health impact.
There are four possible options for the disposal of spent batteries - composting, incineration, land filling or recycling. Composting is obviously not intentionally utilized as most battery systems are simply not biodegradable. Incineration likewise is not a preferred option because of the low calorific value of batteries. They simply do not burn well, and their mass is not substantially reduced by the incineration process. However, incineration is utilized in some countries where land filling is not as viable an option to reduce volumes of combustible wastes. In Japan and some European nations which have little or no available landfill space, incineration of municipal solid waste (MSW) has become a necessity. Batteries which are invariably contained in municipal solid waste will not be reduced in volume by incineration and will most likely partition to the clinker ash or residue from the MSW incineration process. In some cases, small consumer batteries may be broken apart, battery materials oxidized or volatilized, and subsequently recondensed on the fine fly ash from the incinerator. Air emission pollution control devices should capture better than 99% of these fly ash emissions (Chandler 1995), but then the fly ash must generally be subsequently landfilled. All in all, however, incineration is not particularly well suited for the disposal of batteries, although it must be realized that incineration of the small consumer cells will invariably occur in some countries which utilize incineration for a large share of their municipal solid waste disposal.
If, in fact, toxic or hazardous materials from batteries do concentrate in the fly ash from incinerators and that fly ash is captured by air emission control devices, then that ash must be disposed of as a hazardous waste in landfills. Ultimately what might be required is the derivation of a statistical probability of a specific chemical release of a specific concentration during a specific time period from the landfilled fly ash. There are, for example, provisional tolerable daily or weekly intakes (PTWIs) for certain materials established by the World Health Organization (WHO) which well might be used to limit the amounts of certain battery metals from land filling. Total life cycle impact analyses may be utilized to help establish those limits. However, it should also be mentioned that the WHO tolerable daily intake levels for cadmium range from 70 |ig per day for the average 70-kg man to 60 (ig per day for the average 60-kg woman.
Cadmium daily intake levels in most OECD nations have been decreasing steadily since the 1970s and today range from 10 to 20 fig per day, well below any levels of human health concern (International Cadmium Association 1999). These relationships are shown in Figure 6.
World Health Organization Daily Tolerable Intake
19« 1900 1996
Figure 6. Daily Cadmium Intake Levels for General Population
Thus, land filling of incinerator ash from batteries may not be a significant problem and releases through this waste disposal option may not be as great as feared by some. The two most likely options for the disposal of spent batteries today are land filling and recycling. Land filling is currently the most widely used option, as it is the most widely used disposal option for all municipal solid wastes in OECD nations. A recent report (OECD 1998) indicated that an average of 63% of the municipal solid waste in OECD
nations was land filled, an average of 17% was incinerated, and the balance of 20% was recycled or composted. However, even if batteries are land filled, it is by no means certain that this disposal option poses an immediate threat to human health and the environment. For example, a Swiss review by the University of Berne for the OECD (Eggenberger and Waber 1998) on landfill leachate data from landfills in Canada, Denmark, France, Germany, Italy, Japan and Switzerland indicated that the vast majority of leachate samples passed the World Health Organization's (WHO) recommended cadmium drinking water standard of 3 (ig per liter. Some of the data included in this survey were obtained from 50-year old unlined landfills, which theoretically should represent a worst case environmental impact scenario. Thus, the present disposal of NiCd batteries in landfills does not appear to pose an unwarranted risk from the perspective of leaching cadmium into the environment and entering the human food chain.
Even when considered on a long term basis, there is considerable doubt that the presence of land filled battery metals such as lead, zinc, and cadmium would have the catastrophic environmental effects which some have predicted. Studies on 2000-year old Roman artifacts in the United Kingdom (Thornton 1995) have shown that zinc, lead and cadmium diffuse only very short distances in soils, depending on soil type, soil pH and other site-specific factors, even after burial for periods up to 1900 years. Another study in Japan (Oda 1990) examined nickel-cadmium batteries buried in Japanese soils to detect any diffusion of nickel or cadmium from the battery. None has been detected after almost 20 years exposure. Further, it is unclear given the chemical complexation behavior of the metallic ions of many battery metals exactly how they would behave even if metallic ions were released. Some studies have suggested, for example, that both lead and cadmium exhibit a marked tendency to complex in sediments and be unavailable for plant or animal uptake. In addition, plant and animal uptake of metals such as zinc, lead and cadmium has been found to depend very much on the presence of other elements such as iron and on dissolved organic matter (Cook and Morrow 1995). Until these behavior are better understood, it is unjustified to equate the mere presence of a "hazardous" material in a battery with the true risk associated with that battery. Unfortunately, this is exactly the method which has been too often adopted in comparison of battery systems, so that the true risks remain largely obscured.
These caveats notwithstanding, there is still little argument that the most preferred option for the disposal of spent batteries is obviously collection and recycling. Not only does this option greatly reduce any risk which may exist, but it conserves valuable natural resources as well. Today, recycling is viewed as the best human health and environmental option for the disposal of spent batteries, and it is the fastest growing option. Lead acid batteries have already achieved impressive recycling rates, better than 90% in the United States, and growing all over the world. The questions surrounding recycling of NiCd batteries are not whether it is or is not the best disposal option, but only how to improve collection rates, how to finance collection and recycling programs to improve returns, how to label batteries to maximize collection, and how to measure recycling rates. With NiMH and Li-ion batteries, the issues are developing the recycling technologies to improve materials recovery. With the alkaline manganese and carbon zinc batteries, the questions revolve more around the economics of the collection and recovery processes.
Obviously collection and recycling of a spent battery prevents the entry of the majority, probably greater than 98%, of the battery's weight into the environment after use. However, there are other environmental impact factors which also must be considered with regard to recycling. For example, when comparing battery systems, it is instructive to compare the relative energies required to recycle various battery systems. Nickel-iron, nickel-cadmium and lead acid batteries are relatively easy to recycle because the reduction of nickel, iron, cadmium and lead oxides back to their pure metals requires less energy than the reduction of the oxides of other battery metals such as zinc, manganese, chromium, titanium, zirconium, lithium and the rare earth metals which are constituents of alkaline manganese, nickel metal hydride and lithium-ion batteries.
Another factor is the emissions associated with the production of battery metals by the recycling process as opposed to production from virgin ore. There have been many studies to demonstrate that recycling requires far less energy input than production of metal from virgin ore (Gaines 1994), but there are also now studies to indicate that emissions from recycling are lower as well. One report (Geomet Technologies 1993) on electric vehicle NiCd batteries, for example, compares cadmium emissions from production and recycling and finds that recycling emissions are roughly 10 to 100 times lower. These results are summarized in Table VI.
Considered from another point of view, three estimates of the degree of materials recovery from the recycling of NiCd batteries all place that recovery rate at greater than 99%. Similarly high recoveries would be expected for the recycling of nickel-iron and lead acid batteries, but recovery rates from recycling of alkaline manganese, nickel metal hydride and lithium ion batteries might be somewhat lower because of the high
Table VI. Cadmium Emissions from Production and Recycling NiCd Batteries*
Production Emissions Recycling Emissions
(grams Cd per KW-hr) (grams Cd per KW-hrl
Water 0.40 to 2.4 0.0014
Land Negligible Negligible
* Source: Geomet Technologies 1993
energies required and the difficulty of reducing some of the battery metal oxides present in these systems. For example, anywhere from 10% to 20% of the total weight of nickel metal hydride batteries might be lost in the slag during the recycling of these batteries due to the presence of very reactive metals (chromium, aluminum, magnesium, vanadium, zirconium, titanium, rare earth elements) which are strong oxide formers and very difficult to reduce. While it has been suggested that this slag could be utilized for other applications, some environmentalists and regulators argue that such "downgraded" applications do not constitute true recycling. Thus, it is possible to recover a very high percentage of the material in a spent battery, and no doubt recovery technology will improve in the future to allow high degrees of materials recovery from all battery systems. However, the efficiency of the collection process for spent batteries and the efficiency of the metal recovery process are both factors which will affect the overall environmental and human health impacts of battery systems.
Once a complete energy and materials inventory of all of the various steps in a battery's life cycle has been established, the next steps are to categorize the inventory items into various groups. In general, these impacts have been realized on three areas:
• Natural Resources
• Human Health Impacts
• Ecological or Environmental Impacts
Determining the impact assessment requires classification of each impact into one of these categories, characterization of the impact to establish some kind of relationship between the energy or materials input/output and a corresponding natural resource/human health/ecological impact, and finally the evaluation of the actual environmental effects. Many life cycle analyses admit that this last phase involves social, political, ethical, administrative, and financial judgments and that the quantitative analyses obtained in the characterization phase are only instruments by which to justify policy. A truly scientific life cycle analysis would end at the characterization phase, as many of the decisions made beyond that point are qualitative and subjective in nature.
The inventory analysis determines all of the energy and materials inputs in a battery's life cycle and all of the outputs which could have an environmental or human health impact. These outputs include direct emissions from all production and manufacturing processes, including emissions from the energy production processes, and from the use, maintenance, recycling or waste disposal of the battery. All of these emissions must then be considered on a normalized basis by dividing by the total lifetime energy of the battery. The results are total amounts of emissions per kilowatt-hour of energy. If the battery is not recycled, then virtually the entire weight of the spent battery must be considered as being dispersed into the environment, although as discussed previously, the true risk or immediate impact of land filled or incinerated and land filled batteries may be released over an extended period of time and only to a limited degree.
The great controversy in life cycle analyses arises when specific impact assessment values are assigned for each particular material. There are many systems which have been proposed and the impact values vary widely. Strictly speaking, impact values should be very specific for the specific battery material involved. In practice, most systems employ generic categories such as "nickel and its compounds" or "lead and compounds" and employ human health and environmental impact data from surrogate compounds which are usually those which have been most studied in environmental and human health research. Unfortunately, this practice creates a worst case scenario analysis in that the surrogate compounds are almost always the highly soluble species of a metal compound, designed to yield rapid results in clinical tests, but not indicative of the manner in which battery compounds may behave. Thus, for example, cadmium chloride, the highly soluble cadmium compound and one often utilized in environmental and human health research, may be and often is used as the surrogate for all cadmium metal and compounds, whereas the cadmium compounds present in NiCd batteries are the much less soluble cadmium oxide and cadmium hydroxide.
Even worse, in many analyses, impact values appear to be assigned quite subjectively with no justification or methodology specified. Because of this problem, huge variations in environmental impact values exist from one method to another. Essentially, one can obtain any life cycle analysis result one desires simply by arbitrarily selecting artificially high or low environmental impact values. To have any validity at all, a life cycle analysis must be based on environmental and human health impact values which are rooted in quantitative, measurable indices of a material's effect on human, terrestrial or aquatic life. A 1997 comparison (Morrow 1997) compared the normalized life cycle analysis impact values for four rechargeable battery systems utilizing five different impact assessment techniques. Needless to say, the results were very inconsistent except that lead acid batteries consistently fared well because of their high recycling rate. All of the other battery systems ranged over the entire spectrum from relatively benign to the most toxic depending on the environmental impact assumptions chosen.
For example, the five impact assessment evaluation methods reviewed in the 1997 comparison (Morrow 1997) were as follows:
• CML Method - Developed by The Centre for Environmental Science in Leiden, The Netherlands. The effects of water and air emissions of various chemicals on certain general areas such as eutrophication, energy depletion, greenhouse effect, acidification, winter smog, summer smog, heavy metals and carcinogenicity were expressed in terms of potential rather than real effects.
• EPS Method - The Environmental Priority in Product Design method was developed in Sweden by the Swedish Environmental Research Institute and the Swedish Federation of Industries. This system sets a value to a change in the environment through impacts on human health, biological diversity, production, resources and aesthetic values.
• Tellus Method - The Tellus Method is based on control costs of various air pollutants and considers factors such as carcinogenic potency ranking, oral reference dose ranking or a combination thereof.
• Ecoscarcity Method - Defines a relationship for a given country of given area between the critical level of a pollutant set by the limited carrying capacity of the natural environment and the actual anthropogenic emissions of that pollutant. The countries evaluated by the ecoscarcity method are Switzerland, Netherlands, Norway and Sweden.
• U.S. Environmental Protection Agency Method - Based on an analysis technique developed for EPA by the University of Tennessee, this method considers all major human health and environmental effects of the chemicals including persistence and bioaccumulation. It also includes weighting factors for the actual levels of emissions.
These various evaluation schemes produce widely varying results. For example, in rating the metals utilized in various batteries systems, it was generally found that lead, cadmium and mercury consistently were listed as battery metals with the most adverse environmental or human health impacts. However, it was also noted that nickel, cobalt, chromium and even zinc were listed as materials of concern in some systems. Even more remarkable were some of the relative impact assessment values assigned to some battery metals relative to other battery metals. While this variation can be explained to some degree by the different bases used for the techniques, it also clearly indicates that a life cycle evaluation of a battery system will depend to a great extent upon the evaluation system chosen. For example, the relative environmental impact values assigned to six battery metals according to the five different evaluation techniques are summarized in Table VII. These values are all normalized to a maximum value of 100 which is the most adverse environmental impact effect to allow comparison across the five systems.
There is really very little consistency across these environmental impact assessment methods except that the Swedish and Dutch systems rate cadmium the battery metal with the most adverse effects, while the Tellus and Ecoscarcity Methods rate mercury the most adverse battery metal. Zinc, manganese, nickel and even lead have relatively low effects except in the U.S. EPA system, which however is the one system which is most closely tied to actual quantitative assessments of environmental and human health toxicological end points. What is very surprising is the relatively low impact values for mercury in the Swedish and Dutch schemes given the general worldwide concern for mercury.
Table VII. Relative Environmental Impact Values for Battery Metals Utilizing Various Assessment Evaluation Methods
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